Elgar Encyclopedia of Ecological Economics 1802200401, 9781802200409

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Elgar Encyclopedia of Ecological Economics

Emilio Padilla Rosa Jesús Ramos-Martín

Elgar Encyclopedia of Ecological Economics

ELGAR ENCYCLOPEDIAS IN ECONOMICS AND FINANCE Elgar Encyclopedias in Economics and Finance serve as the definitive reference works in the field. The Encyclopedias present a comprehensive guide to a wide variety of subject areas within economics and finance, and form an essential resource for academics, practitioners, and students alike. Each Encyclopedia is edited by one or more leading scholars, internationally recognized as preeminent names within the field. They each include an overarching collection of entries authored by key scholars within the field, which collectively aim to provide a concise and accessible coverage of the essential areas. Equally useful as reference tools or high-level introductions to specific topics, issues, methods and debates, these Encyclopedias represent an invaluable contribution to the field. For a full list of Edward Elgar published titles, including the titles in this series, visit our website at www​.e​-elgar​.com​.

Elgar Encyclopedia of Ecological Economics Edited by

Emilio Padilla Rosa Professor, Department of Applied Economics, Autonomous University of Barcelona, Spain

Jesús Ramos-Martín Professor Serra Húnter, Department of Economics and Economic History, Autonomous University of Barcelona, Spain

ELGAR ENCYCLOPEDIAS IN ECONOMICS AND FINANCE

Cheltenham, UK • Northampton, MA, USA

© Emilio Padilla Rosa and Jesús Ramos-Martín 2023

With the exception of any material published open access under a Creative Commons licence (see www.elgaronline.com), all rights are reserved and no part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical or photocopying, recording, or otherwise without the prior permission of the publisher.

Entry 6, “Boundary Openness over Natural Information” is available for free as Open Access from the individual product page at www.elgaronline.com under a Creative Commons Attribution NonCommercial-NoDerivatives 4.0 International (https://creativecommons.org/ licenses/by-nc-nd/4.0/) license. Published by Edward Elgar Publishing Limited The Lypiatts 15 Lansdown Road Cheltenham Glos GL50 2JA UK Edward Elgar Publishing, Inc. William Pratt House 9 Dewey Court Northampton Massachusetts 01060 USA A catalogue record for this book is available from the British Library Library of Congress Control Number: 2023941868 This book is available electronically in the Economics subject collection http://dx.doi.org/10.4337/9781802200416

EE VS P

ISBN 978 1 80220 040 9 (cased) ISBN 978 1 80220 041 6 (eBook)

Both authors dedicate the book to their families In memoriam of Herman Daly

Contents

11 Coevolution (socio-biophysical coevolution) Miquel A. Gual and Richard B. Norgaard

List of tablesxi List of boxesxii List of contributorsxiii Prefacexvii Agent-based modelling Ivan Savin

1

2 Agroecology Manuel González de Molina

8

1

12 Common property and environmental governance Sergio Villamayor-Tomás

65

70

3 Agrowth Jeroen van den Bergh

14

13 Complex social-ecological systems75 Pedro L. Lomas

4 Anthropocene Jon D. Erickson

21

14 Consumption Doris Fuchs and Inge Røpke

5

Biodiversity conservation Eduardo García-Frapolli

25

6

Bounded openness over natural information Joseph Henry Vogel, María Eugenia Santori-Aymat, Óscar Tomaiconza, Bryan Steven Cortés-Lumbi, and Miguel Fernández-Maldonado

15 Cost shifting, competition and economic structure Clive L. Spash and Amelia Fuselier

32

7

Bounded rationality Stefan Drews

39

8

Carbon taxes Andrea Baranzini and Sylvain Weber

42

9

Circular economy Ignasi Puig Ventosa and Verónica Martínez Sánchez

48

10 Climate change and social justice Éloi Laurent

56

81

88

16 Critical materials Alicia Valero, Guiomar Calvo, and Antonio Valero

95

17 Degrowth Sam Bliss and Giorgos Kallis

98

18 Deliberative ecological economics103 Jasper Kenter

vii

19 Discounting and climate change Cédric Philibert

112

20 Ecofeminisms Corinna Dengler

118

21 Ecological distribution conflicts Joan Martínez-Alier

124

22 Ecological macroeconomics Peter A. Victor

126

viii  Elgar encyclopedia of ecological economics

23 Ecological unequal exchange Mario Pérez-Rincón

133

38 Environmental justice Beatriz Rodríguez-Labajos

24 Economic anthropology Clemens M. Grünbühel

139

25 Economic system José Manuel Naredo

146

39 The environmental Kuznets curve237 David I. Stern

26 Economy as an open system Óscar Carpintero and Jaime Nieto

152

27 Ecosystem services Brigitte L.G. Baptiste 28 Emergy accounting Silvio Viglia and Sergio Ulgiati 29 Energy return on investment: a unifying principle for socio-ecological sustainability Rigo E.M. Melgar and Charles A.S. Hall

40 Environmental limits Erik Gómez-Baggethun

241

41 Environmental stewardship Jennifer Welchman

246

158

42 Environmental tax reform Paul Ekins

248

162

43 Environmental taxation and the double dividend William K. Jaeger

169

30 Energy transition(s) Mar Rubio-Varas

181

31 Entropy Alicia Valero, Antonio Valero, and Guiomar Calvo

188

32 Environmental accounting Maddalena Ripa and Sergio Ulgiati

191

33 The environmental consequences of inequality James K. Boyce

45 Ethics of quantification Andrea Saltelli and Monica Di Fiore

265

46 Fetish, commodity fetishism and ecosystem services270 Nicolas Koso 47 Future generations Richard B. Howarth

200

252

44 Environmentally extended multi-region input–output analysis259 Klaus Hubacek and Kuishuang Feng

273

48 Georgescu-Roegen’s bioeconomics277 Kozo Torasan Mayumi

34 Environmental ethics Joaquín Valdivielso

205

35 Environmental footprints Kai Fang

211

36 Environmental governance Jouni Paavola

217

50 Human appropriation of net primary production (HANPP) Helmut Haberl, Karl-Heinz Erb, and Fridolin Krausmann

223

51 The human ecological footprint William E. Rees

37 Environmental input– output analysis Mònica Serrano

231

49 Green economy Jonathan M. Harris

284

289

298

Contents  ix

52 Incommensurable values Jonathan Aldred

305

68 Natural capital Robert Costanza

394

53 Industrial ecology Anke Schaffartzik

309

397

54 Institutions Arild Vatn

313

69 Nature-based solutions Francesc Baró and Erik Gómez-Baggethun

55 Joint production Johannes Schiller and Stefan Baumgärtner

319

56 Kapp, Karl William Tommaso Luzzati

326

57 Land grabbing Arnim Scheidel

330

72 Peak-Oil Christian Kerschner

58 Land-time budget analysis Clemens M. Grünbühel

336

59 Languages of valuation Christos Zografos

342

73 Political and institutional ecological economics Peter Söderbaum

431

60 The laws of thermodynamics Gabriel A. Lozada

349

74 Population and environment Hernán G. Villarraga

437

61 Material flow accounting Fridolin Krausmann

357

75 Post-normal science Silvio Funtowicz and Jerome R. Ravetz

62 The maximum power principle Mark T. Brown

363

76 The precautionary principle Andy Stirling

440

63 Metabolic flow Mario Giampietro

368

64 Methodological pluralism Richard B. Norgaard

376

77 Production and economic development447 José Manuel Naredo

65 Multi-criteria evaluation Giuseppe Munda

379

66 Multi-Scale Integrated Analysis of Societal and Ecosystem Metabolism (MuSIASEM)385 Mario Giampietro 67 National accounts and macroeconomic indicators Jordi Roca Jusmet

390

70 Nexus approaches in socio-metabolic research Helmut Haberl

403

71 Payments for ecosystem services410 Esteve Corbera and Santiago Izquierdo-Tort

78 Rebound effect and the Jevons paradox Jaume Freire-González

416

425

453

79 Sensitivity analysis Andrea Saltelli, Arnald Puy, and Samuele Lo Piano

460

80 Sensitivity auditing Andrea Saltelli, Samuele Lo Piano, and Arnald Puy

467

81 Social ecological economics Clive L. Spash, Adrien Guisan, and Carlotta Verita

472

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82 Social metabolism Manuel González de Molina

479

83 Spaceship Earth Óscar Carpintero and Jaime Nieto

486

84 Steady-state economics Herman Daly

491

85 Sustainability versus monetary reductionism Peter Söderbaum

496

86 Sustainable development indicators499 Philip Lawn

87 Uncertainty, risk and ignorance Andrea Saltelli and Jerome R. Ravetz

507

88 Uncomfortable knowledge Mario Giampietro

509

89 Unequal caloric exchange Fander Falconí

514

90 Water footprint Cristina Madrid-López

517

Index522

Tables

4.1 The Great Acceleration of the Anthropocene epoch

23

9.1 Indicators used to monitor the circular economy at the EU Member State level

50

10.1 A typology of climate inequality

57

16.1 Materials and elements included in the critical raw material lists for selected countries

96

18.1 Examples of deliberative methods

106

34.1 Principles of climate justice

208

42.1 ETR scenarios and their impacts

249

44.1 Environmentally extended multi-region input–output accounting framework

260

50.1 Overview of estimates of global HANPP given by different authors

293

60.1 An example of microstates, macrostates, and two interpretations of the Entropy Law

352

61.1 Important flows and indicators of economy-wide material flow accounts

359

65.1 Example of an impact matrix

380

69.1 NBS definitions and relevant societal challenges according to IUCN and the European Commission

397

76.1 Key features of a precautionary appraisal process

444

82.1 Metabolic profile of the three metabolic regimes

484

xi

Boxes 6.1 A pesky cat and pet theories

32

6.2 Stepwise institutionalization of “Bounded openness over natural information”

33

46.1 Biodiversity commodified

270

51.1 The structure and function of ecosystems

301

77.1 Definition of a developed country transcending the metaphor of production and the monetary reductionism inherent in GDP

450

80.1 Rules for sensitivity auditing

469

xii

Contributors

Aldred, Jonathan, Emmanuel College, Cambridge, UK

Costanza, Robert, Institute for Global Prosperity, University College London, UK

Baptiste, Brigitte L.G., Universidad EAN, Bogotá, Colombia

Daly, Herman, School of Public Policy, University of Maryland, College Park, USA

Baranzini, Andrea, Haute Ecole de Gestion de Genève, University of Applied Sciences and Arts Western Switzerland, Geneva, Switzerland

Dengler, Corinna, Institute for Multi-Level Governance and Development, Vienna University of Economics and Business, Austria

Baró, Francesc, Department of Geography, Vrije Universiteit Brussels, Belgium; Department of Sociology, Vrije Universiteit Brussels, Belgium; and Institute of Environmental Science and Technology, Autonomous University of Barcelona, Spain

Di Fiore, Monica, Institute for Cognitive Sciences and Technologies, Consiglio Nazionale delle Ricerche, Rome, Italy

Baumgärtner, Stefan, Albert-LudwigsUniversität, Freiburg, Germany Bliss, Sam, Gund Institute for Environment, University of Vermont, Burlington, USA Boyce, James K., Political Economy Research Institute, University of Massachusetts Amherst, USA Brown, Mark T., Department of Environmental Engineering Sciences, Engineering School of Sustainable Infrastructure and Environment, University of Florida, Gainesville, USA Calvo, Guiomar, Centro de Investigación de Recursos y Consumos Energéticos, University of Zaragoza, Spain Carpintero, Óscar, Applied Economics Department and Group of Energy, Economy and System Dynamics (GEEDS). University of Valladolid, Spain Corbera, Esteve, Institute of Environmental Science and Technology and Department of Geography, Autonomous University of Barcelona, Spain; and Institució Catalana de Recerca i Estudis Avançats, Barcelona, Spain Cortés-Lumbi, Bryan Steven, Department of Economics, University of Puerto Rico–Rio Piedras, San Juan, Puerto Rico

Drews, Stefan, Department of Applied Economics, University of Malaga, Spain Ekins, Paul, University College London, UK Erb, Karl-Heinz, Institute of Social Ecology, University of Natural Resources and Life Sciences, Vienna, Austria Erickson, Jon D., Blittersdorf Professor of Sustainability Science and Policy, Rubenstein School of Environment and Natural Resources, University of Vermont, Burlington, USA Falconí, Fander, Department of Development, Environment and Territory, Latin American Faculty of Social Sciences, Quito, Ecuador Fang, Kai, School of Public Affairs, Zhejiang University, Hangzhou, China; and Zhejiang Ecological Civilization Academy, Anji, China Feng, Kuishuang, Department of Geographical Sciences, University of Maryland, College Park, USA Fernández-Maldonado, Miguel, Department of Economics, University of Puerto Rico–Rio Piedras, San Juan, Puerto Rico Freire-González, Jaume, Institute for Economic Analysis, (CSIC), and Barcelona School of Economics, Spain Fuchs, Doris, Institute for Political Science, University of Münster, Germany

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Funtowicz, Silvio, Centre for the Study of the Sciences and the Humanities, University of Bergen, Norway Fuselier, Amelia, Vienna University of Economics and Business, Vienna, Austria García-Frapolli, Eduardo, Instituto de Investigaciones en Ecosistemas y Sustentabilidad, National Autonomous University of Mexico, Morelia, Mexico Giampietro, Mario, Institute of Environmental Science and Technology, Autonomous University of Barcelona, Spain; and ICREA, Barcelona, Spain Gómez-Baggethun, Erik, Department of International Environment and Development Studies (Noragric), Norwegian University of Life Sciences, Ås, Norway; and Norwegian Institute for Nature Research, Trondheim, Norway González de Molina, Manuel, Agroecosystems History Lab, Universidad Pablo de Olavide, Seville, Spain Grünbühel, Clemens M., Australian Centre for International Agricultural Research, Canberra, Australia Gual, Miquel A., Department of Economics, Quantitative Methods and Economic History, Universidad Pablo de Olavide, Sevilla, Spain Guisan, Adrien, Vienna University of Economics and Business, Vienna, Austria Haberl, Helmut, Institute of Social Ecology, University of Natural Resources and Life Sciences, Vienna, Austria Hall, Charles A.S., Department of Environmental and Forest Biology and Division of Environmental Sciences, State University of New York, College of Environmental Science and Forestry, Syracuse, USA Harris, Jonathan M., Global Development Policy Center, Economics in Context Initiative, Boston University, Massachusetts, USA; and Visiting Scholar, Tufts University Global Development and Environment Institute, Medford, Massachusetts, USA Howarth, Richard B., Dartmouth College, Hanover, New Hampshire, USA Hubacek, Klaus, Integrated Research on Energy, Environment, and Society, Energy Sustainability Research Institute Groningen, University of Groningen, The Netherlands

Izquierdo-Tort, Santiago, Instituto de Investigaciones Económicas, National Autonomous University of Mexico, Mexico City, Mexico Jaeger, William K., Department of Applied Economics, Oregon State University, Corvallis, Oregon, USA Kallis, Giorgos, Institució Catalana de Recerca i Estudis Avançats, Barcelona, Spain; and Institute of Environmental Science and Technology, Autonomous University of Barcelona, Spain Kenter, Jasper, Ecologos Research Ltd, Borth, Wales, UK; and Aberystwyth Business School, Aberystwyth University, Aberystwyth, Wales, UK Kerschner, Christian, Department of Sustainability, Governance, and Methods, MODUL University Vienna, Austria; and Department of Environmental Studies, Masaryk University, Brno, Czech Republic Kosoy, Nicolas, Faculty of Agricultural and Environmental Sciences, McGill School of Environment, Montreal, Canada Krausmann, Fridolin, Institute of Social Ecology, University of Natural Resources and Life Sciences, Vienna, Austria Laurent, Éloi, French Economic Observatory and School of International Affairs at Sciences Po, Paris, France; Ponts ParisTech, Paris, France; and Visiting Professor, Stanford University, California, USA Lawn, Philip, Torrens University, Adelaide, Australia Lo Piano, Samuele, School of the Built Environment, University of Reading, UK Lomas, Pedro L., Lomas, Pedro L., Fundación FUHEM, Madrid, Spain Lozada, Gabriel A., Department of Economics, University of Utah, Salt Lake City, USA Luzzati, Tommaso, Department of Economics and Management, University of Pisa, Italy Madrid-López, Cristina, SosteniPra Research Group, Institut de Ciència i Tecnologia Ambientals, Autonomous University of Barcelona, Spain

Contributors  xv

Martínez-Alier, Joan, Institut de Ciència i Tecnologia Ambientals, Autonomous University of Barcelona, Spain

Rees, William E., School of Community and Regional Planning, University of British Columbia, Canada

Martínez Sánchez, Verónica, Fundació ENT, Barcelona, Spain

Ripa, Maddalena, Italian Institute for Environmental Protection and Research (ISPRA), Rome, Italy

Mayumi, Kozo Torasan, Kyoto College of Graduate Studies for Informatics, Kyoto, Japan Melgar, Rigo E.M., Rubenstein School of Environment and Natural Resources, University of Vermont, Burlington, USA; and Gund Institute for Environment, University of Vermont, Burlington, USA Munda, Giuseppe, European Commission, Joint Research Centre, Ispra, Italy Naredo, José Manuel, Universidad Politécnica de Madrid, Spain Nieto, Jaime, Applied Economics Department and Group of Energy, Economy and System Dynamics (GEEDS), University of Valladolid, Spain Norgaard, Richard B., Energy and Resources Group, University of California, Berkeley, USA Paavola, Jouni, School of Earth and Environment, University of Leeds, UK Padilla Rosa, Emilio, Department of Applied Economics, Autonomous University of Barcelona, Spain Pérez-Rincón, Mario, Universidad del Valle–Instituto CINARA (Institute for Research and Development in Water Supply, Environmental Sanitation and Conservation of Water Resources), Cali, Colombia Philibert, Cédric, Energy and Climate Center, French Institute for International Relations, Paris, France Puig Ventosa, Ignasi, Fundació ENT, Barcelona, Spain Puy, Arnald, School of Geography, Earth, and Environmental Sciences, University of Birmingham, UK Ramos-Martín, Jesús, Department of Economics and Economic History, Autonomous University of Barcelona, Spain Ravetz, Jerome R., Institute for Science, Innovation, and Society, University of Oxford, UK

Roca Jusmet, Jordi, Department of Economics, University of Barcelona, Spain Rodríguez-Labajos, Beatriz, Johns Hopkins University–University Pompeu Fabra Public Policy Center, Department of Political and Social Sciences, Pompeu Fabra University, Barcelona, Spain; and Energy and Resources Group, University of California, Berkeley, United States Røpke, Inge, Department of Planning, Aalborg University, Denmark Rubio-Varas, Mar, Institute for Advanced Research in Business and Economics, Universidad Pública de Navarra, Spain Saltelli, Andrea, University Pompeu Fabra, Barcelona School of Management, Barcelona, Spain Santori-Aymat, María Eugenia, Department of Economics, University of Puerto Rico–Rio Piedras, San Juan, Puerto Rico Savin, Ivan, Institute of Environmental Science and Technology, Autonomous University of Barcelona, Spain; and Graduate School of Economics and Management, Ural Federal University, Yekaterinburg, Russian Federation Schaffartzik, Anke, Department of Environmental Sciences and Policy, Central European University, Vienna, Austria Scheidel, Arnim, Institute of Environmental Science and Technology (ICTA-UAB), Spain Schiller, Johannes, Helmholtz-Centre for Environmental Research – UFZ, Leipzig, Germany Serrano, Mònica, Department of Economics, University of Barcelona, Spain Söderbaum, Peter, Professor Emeritus in Ecological Economics, School of Sustainable Development of Society and Technology, Mälardalen University, Västerås, Sweden Spash, Clive L., Vienna University of Economics and Business, Vienna, Austria

xvi  Elgar encyclopedia of ecological economics

Stern, David I., Arndt-Corden Department of Economics, Crawford School of Public Policy, The Australian National University, Canberra, Australia Stirling, Andy, Science Policy Research Unit, University of Sussex, UK Tomaiconza, Óscar, Department of Economics, University of Puerto Rico–Rio Piedras, San Juan, Puerto Rico

Verita, Carlotta, Vienna University of Economics and Business, Vienna, Austria Victor, Peter A., York University, Toronto, Canada Viglia, Silvio, Italian National Agency for New Technologies, Energy and Sustainable Economic Development, Casaccia Research Centre, Rome, Italy

Ulgiati, Sergio, Department of Science and Technology, Parthenope University of Naples, Italy; and School of Environment, Beijing Normal University, Beijing, China

Villamayor-Tomás, Sergio, Institut de Ciència i Tecnologia Ambientals, Autonomous University of Barcelona, Spain; and Ostrom Workshop, Indiana University, Bloomington, USA

Valdivielso, Joaquín, Department of Philosophy and Social Work, Universitat de les Illes Balears, Spain

Villarraga, Hernán G., Grupo de Población y Ambiente, Universidad Regional Amazónica Ikiam, Tena, Ecuador

Valero, Alicia, Centro de Investigación de Recursos y Consumos Energéticos, University of Zaragoza, Spain

Vogel, Joseph Henry, Department of Economics, University of Puerto Rico–Rio Piedras, San Juan, Puerto Rico

Valero, Antonio, Centro de Investigación de Recursos y Consumos Energéticos, University of Zaragoza, Spain

Weber, Sylvain, Haute Ecole de Gestion de Genève, University of Applied Sciences and Arts Western Switzerland, Geneva, Switzerland

van den Bergh, Jeroen, Institute of Environmental Science and Technology, Autonomous University of Barcelona, Spain; Institució Catalana de Recerca i Estudis Avançats, Barcelona, Spain; and School of Business and Economics and Institute for Environmental Studies, Vrije Universiteit Amsterdam, The Netherlands Vatn, Arild, Department of Environment and Development Studies, Norwegian University of Life Sciences, Ås, Norway

Welchman, Jennifer, Department of Philosophy, University of Alberta, Canada Zografos, Christos, Johns Hopkins University–University Pompeu Fabra Public Policy Center, Barcelona, Spain; and Research Group on Health Inequalities, Environment, and Employment Conditions, Department of Political and Social Sciences, University Pompeu Fabra, Barcelona, Spain

Preface

Ecological Economics is an interdisciplinary field that emerged in the 1980s due to the need to integrate approaches from ecology and economics to provide an adequate understanding of the interrelationship between the economy and nature. The field recognizes the biophysical limitations of the economy, which is considered a subsystem of the biosphere, and analyzes their interdependence. Furthermore, it seeks the development of governance and policy tools to pursue ecological sustainability, human well-being, and social justice. Antecedents of the discipline can be found in some 18th- through 20th-century works on topics that were later ignored by mainstream economics. Some of these are, for example, the works of T.R. Malthus on population and poverty, and the works of W.S. Jevons on non-renewable resources. However, the most direct antecedents of the discipline appear with some contributions in the 1960s and 1970s. We can highlight, for instance, the pioneer analyses of K.W. Kapp on social costs and the interrelation between the economy and nature, and the works of N. Georgescu-Roegen on the importance of the entropy law in the economic process and its implications on the limits of economic growth, followed by contributions by authors such as K.E. Boulding or H.E. Daly, who integrated knowledge from ecology and economics in their seminal contributions. The magnitude of the ecological crises generated by economic activity raised an important environmental concern in the ’70s and ’80s, while economics and natural sciences disciplines alone were unable to provide adequate analyses and political responses. This led some ecologists and economists to promote the integration of knowledge from the natural and social sciences, defining the new field and organizing the first meetings of ecological economists with two workshops on “Integrating Ecology and Economics” in 1982 in Saltsjöbaden, Sweden, and in 1987 in Barcelona, Spain. In addition, in 1989, the International Society of Ecological Economics was established and its offi-

cial journal, Ecological Economics, was launched. Several of the authors who participated in these meetings have also contributed to this Encyclopedia. The new discipline was not a new subdiscipline or school of thought of economics but a new transdisciplinary approach to study the interactions between economic subsystems and nature and provide solutions to environmental and social problems derived from these. Moreover, the field cannot be considered a closed body of knowledge, but is rather in constant evolution, and different visions and sensibilities coexist in it, which we have tried to reflect in this collective work. Ecological Economics has significantly expanded in the last three decades. Being an interdisciplinary and transdisciplinary field that is evolving with the contributions of authors from different natural and social sciences (ecologists, economists, biologists, sociologists, engineers, political scientists, etc.), the range of topics of Ecological Economics is quite wide. It includes theoretical, methodological, and empirical contributions, besides having a strong normative focus, as it orients its efforts to improve environmental policy and governance to enhance well-being, fairness, and environmental quality. Making a complete coverage of the topics of relevance for Ecological Economics would be an impossible task, so this Encyclopedia focuses on a selection of the topics that are more neatly identified with the contributions of the field. There are connections, shared interests, and overlaps with other fields, such as Environmental and Natural Resource Economics, Industrial Ecology, or Political Ecology, among others. While the differentiating trait of Ecological Economics is its focus on the interdependence between economic subsystems and the natural environment in which they are embedded, we have, anyway, not avoided the overlapping with these other fields in the selection of entries, as this overlap is also a trait of the field. The Encyclopedia presents an overview of the development of the discipline, from

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the fundamental contributions of the pioneer authors to the more recent methodological debates in the field. Its content is oriented to academics and students with different backgrounds interested in environmental problems and their solutions from a holistic perspective, to practitioners with responsibilities in the management of natural and environmental resources, and to environmental activists. The Encyclopedia is intended to be a work of reference for anyone— from students and researchers in the field, to colleagues from other fields, social actors, and policy makers—who may be interested in authoritative entry points into a particular concept, debate, or development within Ecological Economics. The entries of the Encyclopedia reflect the relevant state of research, its foundations, key thinkers, and past and present major debates about that particular concept, theory, actor, or issue at hand. The Encyclopedia is integrative and includes the different views and

sensibilities within the discipline in its different entries instead of showing the editors’ own view. We find that the selection of entries covers the main aspects of Ecological Economics and presents an up-to-date overview of the discipline. Given the heterogeneity of entries, as well as the plurality of methods and views in the discipline, we have not imposed a unique structure for all chapters, but they vary according to the different concepts dealt with and to author preference. For this reason, we have also preferred to alphabetically classify the entries, which may make their consultation easier, instead of forcing another structured classification. The Encyclopedia gathers a diverse international group of academics and professionals to whom we are very grateful. We are also much indebted to all the reviewers of the entries, and to Daniel Mather at Edward Elgar Publishing for suggesting this project to us. Emilio Padilla Rosa and Jesús Ramos-Martín

1. Agent-based modelling

changes in particular assumptions or parameters will affect the model’s overall behavior. Applications of ABM can be found now in many different research fields, including ecology, economics, and the social sciences (Niazi and Hussain, 2011).1

Definition

Agent-based models (ABMs), also known as “agent-based complex systems,” “agent-based computational economics,” “complex adaptive systems,” and “individual based models,” are computational models that are used for simulating the behavior of autonomous agents (both individual and collective entities, such as firms or entire countries) and their interaction to assess how the system composed of many agents performs in the aggregate and what governs its outcomes (Macy and Willer, 2002). This bottom-up perspective on systems, be it markets or regions, is a distinct feature of ABMs (Bonabeau, 2002). ABMs are uniquely positioned to describe individual agents as boundedly rational and as learning either from their own experience (individual learning through experimentation or trial and error) or from the peers they interact with (social learning). Sometimes, these two types of learning are combined (Savin and Pushkarev, 2015). In the social sciences, learning through interaction is used more often as it is faster and easier to program. However, as Vriend (1998) has shown, the results of social learning are sensitive to the type of interaction agents have and should therefore be interpreted carefully. Most ABMs are composed of:

Motivation for ABMs

ABMs signify a step away from assumptions of the representative agent that characterize traditional economic policy models that have been exceedingly criticized during the recent global economic recession. In his speech in November 2010, former European Central Bank president Jean-Claude Trichet admitted that “[A]s a policymaker during the crisis, I found the available models of limited help. In fact, I would go further: in the face of the crisis, we felt abandoned by conventional tools.” Three comments are in order here. First, traditional economic models are commonly built around the concept of equilibrium where markets clear and the economy arrives at the unique equilibrium smoothly and with certainty. In contrast, history is full of severe recessions with long periods of recovery (Farmer et al., 2015). Second, in contrast to the representative agent assumption common in economic models, empirical research demonstrates persistent heterogeneity among households in terms of income and consumption choices, as well as among firms in terms of size, growth rate, and productivity (Heckman, 2001). Third, traditional models tend to rely on isolated, rational agents and impose unrealistic cognitive capabilities on them (De Grauwe, 2011), and ignoring direct (sometimes called “local”) interactions between agents in a population of people or organizations (Powell, 1998), thus oversimplifying the process of accessing and processing information. ABMs are designed to relax these restrictive assumptions: they are built to describe the process of forming an aggregate outcome potentially focusing on the out-of-equilibrium analysis, they can accommodate multiple heterogeneous characteristics of its constituent agents, typically represent agent behavior as a set of heuristic rules, and explicitly model agent interaction and learning through social or spatial networks. All these features make ABMs more “realistic,” which is considered to be their main advantage.

1. multiple agents (potentially belonging to several types, like firms or households); 2. their decision-making heuristics (what actions to do under which conditions); 3. learning rules (i.e., the possibility to revise heuristic rules); 4. an interaction topology that could be represented by a social network or a market where agents occasionally meet to interact; and 5. an environment in the form of ecological or economic conditions, such as pollution, scarcity of resources, or inflation. ABMs are typically implemented as computer code, either as custom software or via ABM packages and platforms (Abar et al., 2017; Foramitti, 2021; Gill et al., 2021). Within such a code, one can then explore how 1

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Unique features

Let us discuss some of these ABM properties in more detail. The possibility to account for boundedly rational behavior of agents means that the assumptions of rationality and perfect foresight can be relaxed. Traditionally, economic models are based on rational choice theory with self-interested individuals, who have egoistic preferences and unlimited cognitive capabilities to estimate all possible actions. One of the revolutionary insights over the last few decades, credited with several Nobel prizes, is that actual human behavior is highly sensitive to the framing and context of the decision (Simon, 1956; Tversky and Kahneman, 1974; Thaler and Sunstein, 2008) and exhibits cognitive biases, challenging the application of rational choice theory as an accurate description of individual decision-making. Biases are the result of the limited time people have to gather information and prepare decisions, giving way to routine-based (heuristic) decision-making or emotions influencing decisions. According to Simon (1956), the founder of the notion of bounded rationality, human decision-making is constrained by the cognitive capabilities to process information as well as by the availability of this information. Instead of finding an optimal solution (the absolutely best option among those possible), boundedly rational agents try to find a “satisfying” option in a situation, given their limits on information gathering and their cognitive capabilities. To do so, agents follow heuristics to judge the information they have and choose actions that lead to the more preferred outcome. Nowadays, most ABMs contain some form of boundedly rational behavior (for a review, see Castro et al., 2020). This ranges from limited awareness and myopia to imperfect expectations and mark-up pricing. Conventionally, ABMs employ heuristics as their building block by describing agents as possessing a (usually fixed) set of behavioral rules and, potentially, allowing them to switch between these based on their relative performance. That is, an agent can apply one of those rules for some time, but then compare performance with their peers, and subsequently imitate the more successful strategies (i.e., use social learning; d’Andria and Savin, 2018).

Ivan Savin

Furthermore, typical for ABMs is local interaction between agents. ABMs are not bound to follow simplified rules of random interaction, where essentially everyone can interact with everyone with equal probability. Instead, similar to the real world, agents have different propensities to interact depending on their location in the spatial or social network (Breschi and Lissoni, 2009; Savin and Egbetokun, 2016; van den Bergh et al., 2019). Importantly, one can tune the network of interest to empirical data (such as spatial location of households, location of power plants in the electricity grid, or network of interbank loans) to make analyses more accurate. In particular, one can estimate the topology and density of a real social (Rand et al., 2015) or technological (Korzinov and Savin, 2017) network, or use data on geophysical location (Crooks and Castle, 2012) to derive network parameters for the model. Due to their numerical nature, ABMs can easily measure the effect of distinct network topologies, be it regular lattice, small world, or scale-free networks to show what network type is more suitable for rapid propagation of information or diffusion of a social norm (Konc and Savin, 2019). Finally, ABMs uniquely can combine market and social or non-market interactions, allowing users to connect distinct research and policy areas. For example, this allows the systematic study of combinations of marketand non-market-based policy instruments, such as taxes and information provision. A more subtle issue can be addressed then as well—namely, how market choices and social interactions influence one another and, allied to this, how associated policies interact (i.e., whether they are synergistic or antagonistic). For example, a recent ABM study by Konc et al. (2021) discusses how social networks can reinforce environmental and climate policies of a regulatory or pricing nature. In particular, the authors study how peer pressure among consumers can magnify the effectiveness of carbon pricing. To this end, they build a model of consumption decisions driven by socially embedded preferences formed under the influence of peers in a social network. Konc et al. find the social multiplier of environmental policy to equal 1.3, allowing the effective tax to be reduced by 38 percent.

Agent-based modelling  3

Guide to developing ABMs

Frenken (2006) presents four criteria of “good” ABMs. First, ABMs usually use “stylized facts,” such as widespread empirical regulations. If a model replicates a large number of stylized facts, it is considered to be good under the ROSF (Replication Of Stylized Facts) criterion. Second, ROAS (Realism Of the ASsumptions) means that a considered model should not contradict well-known behavioral or environmental facts. Note that traditional neoclassical models use the opposite philosophy, according to which the prime quality criterion of models is the ability to make accurate predictions. Third, KISS (Keep It Short and Simple) implies that, all else equal, a simpler model that can explain a certain phenomenon should be preferred to a complex model. Therefore, one should focus on the main actors and interdependencies, avoiding unnecessary complexity. One could in fact apply this criterion to any model, and not just ABMs. The peculiarity of ABMs is that, due to their algorithmic nature being written as a computer code, one can more easily add complexity to them and still obtain the results. The situation is different with traditional general equilibrium models, as further assumptions make them harder to solve, so there is an incentive to keep them simple. Therefore, the primary motivation for ABMs to stay simple is the clarity about the cause–effect nexus. Finally, TAPAS (Take A Previous model and Add Something) states that incremental approach is usually the most appropriate in developing ABMs. This is because, using existing models, we contribute to the accumulation of knowledge and minimize the time for development of the new model, making new models at the same time more comparable in contrast to models built from scratch. This does not mean that there is no place for experimenting with new models, but that one should take advantage of existing models, which make new models faster to construct and easier to understand.

Possible classifications

There are different approaches to classifying ABMs, and most are concentrated around the model’s scope of analysis, its purpose, and the resulting ABM complexity. Gerst

et al. (2013) propose four classes of ABMs based on the level of detail devoted to techno-economic analysis and the scope of countries/regions involved: a single market concentrating on diffusion of a given technology; competition between several technologies within a market; considering underlying interdependencies of actors, markets, flows of goods, energy, money, and pollution within an entire country; and international policy and country interaction. A different taxonomy has been proposed by Boero and Squazzoni (2005). Their idea is to classify ABMs based on their target and empirical relation, distinguishing between “case-based models,” “typifications,” and “theoretical abstractions.” Case-based models focus on a specific time and space phenomenon, such as the evolution of a certain industry in a given country over a pre-specified period (these models are also frequently called “history friendly” ABMs; see Garavaglia, 2010). Their main purpose to identify the main factors explaining the specificity of the case. Case-based models are usually rich in detail and empirical data (Janssen and Ostrom, 2006) focusing on replicating a particular historical trend and predicting its further evolution, such as the diffusion of renewable energy in Germany between 1990 and 2010 (Herrmann and Savin, 2017). Typifications, in contrast, represent models that replicate features applicable to a wide range of empirical cases (industries, countries, periods of time). Instead of replicating exact time series observed in real words, typifications concentrate on the stylized facts and try to explain mechanisms generating those across a wide range of empirical contexts and, thus, claiming generality of those mechanisms. An example of a typification model would be a model of herding behavior by De Grauwe (2011). Finally, theoretical abstractions are the most simplistic and general among the three types, representing a metaphor of social reality rather than its accurate description. Instead of detailed economic insights, theoretical abstractions aim to provide intuition for debates about, for example, emerging social order (the Schelling, 1971, model) or firm motives to form research and development alliances (Savin, 2021).

Ivan Savin

4  Elgar encyclopedia of ecological economics

Applications of ABMs, notably in ecological economics

Since ABMs by construction serve as an experimentation laboratory where one can control the factors in question and alternate their magnitude (Sun and Tesfatsion, 2007), they are well suited to: 1. understand how stable economic systems are behaving under alternatively specified conditions; 2. explain why certain empirical regularities emerge and persist over time; and 3. evaluate alternative policies in terms of environmentally or socially desirable outcomes. In line with this, ABMs have been applied to a variety of research questions. This means there are good lessons to learn from past experience. In particular, ABMs have been applied to study housing markets (Geanakoplos et al., 2012), automobile markets (de Haan et al., 2009), financial markets (LeBaron, 2001), labor markets (Dawid and Gemkow, 2014), diffusion of knowledge (Cowan and Jonard, 2004), and opinions (Weisbuch et al., 2002). There is also a class of large-scale agent-based macroeconomic models representing complex macro-evolutionary interactions between financial, energy, capital, and consumption good markets (Dosi et al., 2010; Dawid et al., 2012; Ponta et al., 2016). These macroeconomic studies are designed to examine systemic policy effects demonstrating, for example, differential impacts that fiscal and monetary policies have for countries with different levels of income inequality (Russo et al., 2015), and how this income inequality puts the stability of the financial sector in question (Safarzynska and van den Bergh, 2016). Related to ecological economics, ABMs have been applied to study, among others, electricity (Sun and Tesfatsion, 2007) and emissions trading markets (Foramitti et al., 2021a, 2021b), diffusion of eco-innovations (Bleda and Valente, 2009), and regulation of residential electricity consumption (Lin et al., 2016; Rai and Henry, 2016). Classifying 61 ABM studies devoted to climate-energy policy, Castro et al. (2020) demonstrate that 25 were devoted to emission reduction,

Ivan Savin

another 21 to product/technology diffusion, and another 15 to energy conservation. Furthermore, due to their ability to exploit agent heterogeneity and realistic geographical environments with the explicit spatial location of agents therein, ABMs are frequently used in (urban) land use and water planning models (see Huang et al., 2013, and Lindkvist et al., 2020, for reviews). For example, Gawith et al. (2020) study farmers’ behavior in response to climate change, Müller-Hansen et al. (2019) explore possible actions of ranchers to reduce deforestation, while Carrella et al. (2020) fit an ABM of boundedly rational behavior to predict decisions of fishing vessels. These applications belong to the literature on common pool resources addressing the conditions necessary for agents to change their behavior to avoid the collapse of a resource in the longer term (Ostrom et al., 1994). More recently, ABMs have been used to analyze climate policy acceptance and feasibility. While it is clear that policy implementation depends on social support, few studies have actually examined their co-evolution. An ABM by Gerst et al. (2013) describes countries negotiating targets for international climate policy and taking preferences of firms and households into account. Isley et al. (2015) design an ABM to explore the role of industrial lobbyists on the stringency of climate policies. Konc et al. (2022) compare carbon tax and performance standards with different stringency trajectories over time and allow voters to veto policy implementation if a majority is against it. The study finds that, by redistributing carbon tax revenue to low-income households and gradually increasing the tax rate, one can achieve maximum public support.

Acknowledgment

I am grateful to Jeroen van den Bergh for helpful feedback. The usual disclaimer applies. Ivan Savin

Note 1.

For reviews of ABM applications to environmental issues, see Brown et al. (2017) and Groeneveld et al. (2017), and to climate policy issues, see Castro et al. (2020) and Savin et al. (2023).

Agent-based modelling  5

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Ivan Savin

2. Agroecology

gled against the Green Revolution and agricultural industrialisation. A recent issue of Agroecology and Sustainable Food Systems, coordinated by Miguel Altieri and Clara Nicholls (2017), includes multiple examples of this close connection. The pro-campesino movement and alternatives to the corporate food regime have been part of agroecology’s DNA since its inception. They have constituted agroecology’s most notable mark of identity, together with political ecology and popular environmentalism.

Introduction

Food consumption patterns have become a major cause of global unsustainability in terms of both human health and the health of agroecosystems. Productive imbalances between countries, unequal distribution of land, and control of world agricultural markets by large agri-food corporations and financial institutions have turned food insecurity, endemic hunger, and widespread poverty into structural characteristics of the global agricultural system. Such a disequilibrium reflects the exhaustion of the technological model of intensive agriculture, which emerged after the Second World War and gave rise to the so-called Green Revolution. The fact that this model is unviable is increasingly recognised by academics, research centres, international organisations, and even government agencies. New alternative models have been advanced. Among them, agroecology has emerged precisely to address this issue and agroecological implementations have achieved promising results. Agroecology goes back a long way, though the term cropped up in the 1970s to study phenomena such as how weeds and pests interact with cultivated plants. The meaning of the term gradually expanded: it now refers to a conception of agricultural activity that is more integrated in the environment and more socially balanced. In 1928, in the field of agronomy, Klages (1928) raised the need to consider the physical and agronomic factors that influenced certain crop species’ adaptations. Yet it was not until the 1970s that the close links between agronomy and crop ecology were set forth (Hecht, 1995). And it was not until the early ’80s that social aspects were introduced as key explanatory variables, especially in the analysis and design of rural development programmes (Buttel, 1980). The story is fairly well known, and a summary can be found in Guzmán et al. (2000) or Wezel et al. (2009). Agroecology did not emerge, however, merely as one more current – resulting from scientific developments – within the discipline of agronomy or other agricultural sciences. It was a consequence of the theoretical and, above all, practical demands of peasant movements, especially in Latin America, as they strug-

Agroecology as a scientific practice

Agroecology conducts science a little differently from scientific-technical rationality. Its practice is rooted in the ecological paradigm and constitutes an epistemological alternative to conventional science’s approach to agricultural activity. Faced with the fragmentation of knowledge, agroecology relates social and economic aspects with environmental or agronomic ones. That is why agroecosystems are understood as an expression of social and ecological relations (Guzmán & González de Molina, 2017). Agroecology questions the social usefulness of the knowledge it produces. In this way, its quality does not derive from measurements made by isolated scientists and according to the scientific rationale itself, but also from society’s participatory evaluation. This is agroecology’s normal way of operating which, combined with social movements, seeks solutions to the food system’s current crisis. Agroecology thus practices ‘science with the people’, without whom knowledge cannot be constructed; the latter, therefore, conditions its methods, including that of participatory action research (Méndez et al., 2013; Guzmán et al., 2013). Agroecologists firmly support interdisciplinary dialogue, enabling encounters between different fields of knowledge and worldviews (Freire, 1976; Toledo & Barrera-Bassols, 2008). Scientific knowledge and local knowledge, especially that of peasants, meet and amalgamate to collectively build or co-create agroecological knowledge. Moreover, agroecology is a ‘hybrid discipline’ (Toledo, 1999) that can be integrated into sustainability science. Its approach is fundamentally different from that of conventional agricultural sciences: rather than seeking maximum production, agroecology is engaged in a quest for sustainability. It is eth8

Agroecology  9

ically committed to altering the social, economic, and institutional conditions that have led to the current state of unsustainability. Its determination is unbreakable. Its ethical obligation therefore resides in the realisation of a social transformation, in the imperative to act, both to implement the agroecological transition (the construction of an alternative regime), and to fight against the status quo that has led to the ecological crisis. At the very heart of agroecology lies a commitment to change and action.

Agroecology as a productive practice

Agroecology’s main objective is to achieve the sustainability of agroecosystems and of the food systems in which they are integrated. Far from a conventional vision, however, it defends a conception that is at once multidimensional and integrated: an alternative way of understanding the environmental health of agroecosystems, the economic viability of food activity, social equality – considering it a determining factor – and even the radical defence of participatory democracy and food sovereignty as guarantees of food sustainability. The latter, one must add, is not always considered in politics. Further still, agroecology asserts the close links between the four dimensions mentioned above. Agroecosystems are ecosystems that humans manipulate and render artificial in order to capture and convert solar energy in the form of biomass that can be used as food, medicine, fibre, or fuel (Altieri, 1995). From a thermodynamic standpoint, agroecosystems are complex adaptive systems that dissipate energy to counteract the law of entropy. Unlike ecosystems, they are unstable and require external energy, materials, and information (Gliessman, 1998). These flows are channelled through agricultural works directed towards the production of biomass, ensuring repetition through successive cultivation or breeding cycles, interfering in carbon, nutrient, and hydrological cycles as well as biotic regulation mechanisms. In peasant-managed agroecosystems, the additional energy and material input comes from biological sources: human labour and animal labour, maintaining a strict territorial dependence. In industrially managed agroecosystems, the energy and additional materials also

come from the direct and indirect use of fossil fuels and metallic and non-metallic minerals. In such systems, most of the energy generated as biomass is driven out of the system, both in the form of food or fibre and harvest waste. These agroecosystems are mere ‘energy transporters’ and can hardly be considered sustainable (Gliessman et al., 2007, 17). Agroecology therefore understands agricultural sustainability as an isomorphic property of ecosystems: the more agroecosystems resemble natural ecosystems in their organisation and functioning, the greater their sustainability. In this way, an agroecosystem’s sustainability would be positively correlated with the quantity and quality of its internal loops and thus with its internal energy flows that circulate to reproduce the fund elements (Guzmán & González de Molina, 2017). Both criteria are satisfactorily met if the biogeochemical cycles are completed at a landscape scale and it is not necessary to use external inputs, maintaining the market’s autonomy and independence. Market autonomy and independence are also key factors to recover agroecosystems’ economic viability (Ploeg et al., 2019). Indeed, the sustainability of agroecosystems depends on their level of biodiversity, the maintenance of fertile soil, and so on. This implies that a share of the generated biomass must recirculate to meet both productive and reproductive functions, which are essential to the agroecosystem itself: seeds, animal labour, organic matter in the soil, functional biodiversity, and the like. For example, only biomass can nourish the food chains that sustain both the agroecosystem’s edaphic life and biodiversity generally. The deterioration of the colonised or appropriate territory cannot be compensated with energy and external materials or that of a nature other than plant biomass. In this way, agricultural industrialisation can be understood as a process in which dissipative structures of a biophysical nature, present in agroecosystems and maintained by peasants through integrated management, are replaced by human-made dissipative structures – or, in economic terms, by technical means of production achieved through the market and, to a lesser extent, via state intervention. In this sense, it is worth recalling here the idea, expressed by ecological economists, that natural capital cannot be replaced by Manuel González de Molina

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manufactured capital, just as not all types of energy are interchangeable or have the same usage (Giampietro et al., 2015). The destruction of internal loops has degraded agroecosystems’ major fund elements (water, soil, climate, and biodiversity) in such a way that these elements offer fewer and lower-quality ecosystem services. The latter reflects the severe environmental impacts caused by the industrial agriculture model. Only a sustainable management of agroecosystems can guarantee their optimal performance. Consistent with the above, agroecosystem functional autonomy and independence from external inputs is not only a guarantee of sustainability; they also mark the difference between agroecological management and all other proposed forms of management, including those that include the term ‘sustainable’ in their objectives (climate-smart agriculture, precision farming, sustainable intensification, etc.). This same biomimetic approach uses agroecology when addressing the food system’s organisation. Agroecology pursues the expansion and consolidation of shorter and more sustainable distribution and marketing channels, promoting territorial or local food systems. This link with the territory is fundamental, not only because it seeks a maximum coupling between food and agricultural production locally, but also because the territory gives meaning and provides identity and cultural significance to the act of eating, facilitating agroecosystem anchoring. The chain’s territorial approach favours, therefore, the location of agri-food processing activities in areas close to farms; the grouping of producers to sell in common, organise production, and regulate – as well as ensure – supply; and, of course, the establishment of the necessary logistic infrastructures. It also allows an effective coordination between production and consumption, as well as alliances with other non-food actors present in the territory, making it possible to anchor agroecological innovations through stable transformations of local food regimes. Finally, this territorial perspective facilitates alterations to the most characteristic consumption patterns that sustain the current regime: the rooting in food traditions encourages a transition towards a healthier diet, with less processed food and less animal protein, a diet that is based more on fresh consumption and seasonal products Manuel González de Molina

than on highly processed foods that originate from distant lands and are excessively costly in energy terms.

The importance of economic viability

Agroecology’s understanding of economic viability is not limited to employment or higher agricultural-product prices: farmers and peasants should receive a level of income high enough to enable them to live with dignity and to meet – according to the country – an average family’s expenses (Ploeg et al., 2019; González de Molina et al., 2019). This implies challenging the industrial agricultural production model, which does not provide sufficient income, destroys jobs, eliminates the peasants themselves, or subordinates them to the logic of large food corporations, turning them into a source of capital accumulation and cheap raw materials. The vicious circle of ever decreasing prices received by farmers, which they attempt to compensate by using increasingly expensive and inefficient inputs, has turned agricultural activity into a continuous source of benefits for the input industry, large agri-food companies, and big retail. Some technological models have been advanced as an alternative to the industrial agriculture model in crisis (Climate-Smart Agriculture, Precision Farming, Sustainable Intensification, etc.). In reality, however, such models do not promote change; rather, they seek to maintain the system by increasing dependence on commercial technologies. The same applies to the new generation of agricultural technologies (robotics, drones, gene editing, big data, blockchain, etc.) that promise substantial agricultural production improvements. These new technologies are expensive and only accessible to large farms and capital-intensive agricultural enterprises. In any event, these technologies, rightly called hyper-technologies, are far beyond the control of producers, especially family members, and thus strengthen the rule exercised by large corporations. These technologies reinforce the model in which agriculture has become a vast market for the input industry and where consumers are passive buyers of food products that are in the hands of agribusiness and big retail.

Agroecology  11

The importance of social aspects in agricultural sustainability

Agroecology cannot look the other way or coexist with hunger, malnutrition, or poverty, nor with the accumulation of wealth or the exploitation of others by a few, be they social groups or countries. This conclusion is not only a moral imperative, it derives from the observation of the dynamics of agroecosystems (Vandermeer, 2009). Constantly underlying the degradation of agroecosystems are social and economic imbalances: they have pushed towards productive intensification or specialisation beyond their own physical limits. The industrial agriculture model undermines the social base that supports the environmental system, but not only that: its capacity to provide a dignified life for farmers has been declining alarmingly. The prices of farmers’ products have continued to decline as a result of economic policies and dominant valuation systems. Farmers have fought against this downward trend in prices received by producing more on their farms or saving costs (i.e. eliminating labour). The consequence has been rising inputs and increased market dependence. These inputs contain comparatively more added value than agricultural products, as they are manufactured and run using fossil fuels, so their purchase price has also risen steadily. The result has been a falling income and systematic employment destruction. The struggle for social equality along the agri-food chain is a commitment that goes hand in hand with agricultural sustainability. Without social equality, there is no sustainability (in its twofold social and gender dimension): under capitalism, which is based on maximum inequality (and, therefore, high entropy), there is no possible equity and, therefore, sustainability. Agroecology and capitalism are, hence, not compatible. Capitalist accumulation, which underpins extended reproduction, generates, per se, social inequality and high resource consumption. Equality, from an agroecological perspective, signifies much more than a merely cosmetic egalitarianism. Within the agricultural (production) sector, it involves equal access to land and water (agrarian reform). And outside the sector, it implies an equal treatment with other sectors and valuation systems that bring about a fairer remuneration of agricultural goods and services. This

entails a redistribution of power in the food chain.

The importance of sustainability’s political dimension

In recent years, agroecological reflections on politics and institutions have deepened, and a new field has emerged, political agroecology, which attempts to systematise the ideas on these issues (Giraldo, 2018; González de Molina et al., 2020). The current institutional framework regulates agri-food markets for the benefit of conventional production, the big interests of the input industry, large agro-industrial companies, and large retail to the detriment of consumers, producers, the environment, and health. It is agroecology’s mission to design institutions that facilitate the changes and adaptations necessary to shift from the corporate regime to a new, sustainability-based regime. This transition is not only economic or technological in nature, but also political; therefore, the shift is not only local, but global. It implies changing material production systems but also modifying how food-related public goods are reproduced and socially distributed. These objectives are unfeasible without change of a political nature. Public policies must reverse the current situation by introducing measures and regulations that remould the monetary and fiscal incentives enjoyed today by conventional producers and consumers and that so harm organic production. But for this, it is necessary to exert pressure, as the large food corporations do, imposing a new kind of institutionality, preferably through social mobilisation. Agroecology calls for a democratic conception of food sovereignty, understood not only as the right to food, but also as the democratic right to decide on what products are eaten and how they are produced, distributed, and consumed. Agroecology also claims that the dimensions of sustainability interact among themselves and that an agroecosystem or food system’s sustainability (or unsustainability) is the interconnected result of the four dimensions. All of them must maintain a congruent relationship. One of these dimensions is often a limiting factor of sustainability. In accordance with agroecosystems’ socio-ecological nature, the producers who manage the farms and the landscape in which they are located Manuel González de Molina

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play that limiting role. They constitute the key element of any socio-ecological relationship and are the source of the information flows on which the agroecosystems’ management is based. In this sense, the dynamics of agroecosystems are explained by the reproduction vicissitudes of families or domestic groups that contribute agricultural work. The major source of the encountered problems has been unequal access to basic goods and services or the unfair distribution of agricultural income. In other words, inequality in the access and enjoyment of natural resources pushes towards greater metabolic efforts than those that would be necessary if such a distribution were more egalitarian. Consequently, from an agroecological viewpoint, social equality is the key to sustainability.

Agroecology as a collective practice

Agroecology implies a strong commitment towards the transformation of realities. This commitment is twofold: i) a practical commitment, through changes in the management of agroecosystems and the agri-food system as a whole; and ii) a commitment to collective action to generalise change and achieve sustainability. This leads to a close connection with social movements. No scientific practice can avoid direct participation in the agroecological transition. Agroecology cannot be compartmentalised into different practices (production, science, and social movement or engagement): there is only one agroecological practice because no dimension can work in isolation. Agroecology’s priority is to become an alternative to the corporate regime. This is only possible through its scaling out, but above all, its scaling up. Agroecology does not make sense as a minority current or one confined to small social spaces or groups. In addition to the mobilising of peasants and the new peasantry, especially women, agroecology also calls for consumers to take action, making citizens’ healthy eating the epicentre of demands for sustainable practices along the entire food chain, from production and distribution to consumption. Imposing institutional changes is not, however, an easy task. Lobbies, formed by large input, distribution, and food agribusiness companies must be confronted, together with the political organisations that support them. This goal is only Manuel González de Molina

attainable through social mobilisation as well as demands for regulations and public policies that make change possible. Politicising consumption is the most effective way to build majorities of change around agroecology’s main goal: an alternative food regime. Manuel González de Molina

References

Altieri, M.A. (1995), Agroecology: the science of sustainable agriculture, 2nd ed. Boulder, CO: Westview Press. https://​doi​.org/​10​.1201/​ 9780429495465 Altieri, M., & Nicholls, C.I. (2017), Agroecology: a brief account of its origins and currents of thought in Latin America. Agroecology and Sustainable Food Systems, 41(3–4): 231–7. https://​doi​.org/​10​.1080/​21683565​.2017​ .1287147 Buttel, F. (1980), Agriculture, environment and social change: some emergent issues. In F. Buttel & H. Newby (eds.), The rural sociology of advanced societies. Montclair, NJ: Allenheld, Osmun and Co., pp. 453–88. Freire, P. (1976), Education, the practice of freedom. London: Writers & Readers Publishing Cooperative. Giampietro, M., Mayumi, K., & Sorman, A. (2015), Energy analysis for a sustainable future: multi-scale integrated analysis of societal and ecosystem metabolism. London: Routledge. Giraldo, O.F. (2018), Ecología política de la agricultura. Agroecología y posdesarrollo [Political ecology of agriculture. Agroecology and post-development]. San Cristóbal de Las Casas, Chiapas, México: El Colegio de la Frontera Sur. Gliessman, S.R. (1998), Agroecology: ecological processes in sustainable agriculture. Ann Arbor, MI: Ann Arbor Press. Gliessman, S.R., Rosado-May, F., Guadarrama-Zugasti, C., Jedlicka, J., Cohn, A., Méndez, V.E., Cohen, R., Trujillo, L., Bacon, C., & Jaffe, R. (2007), Agroecología: promoviendo una transición hacia la sostenibilidad [Agroecology: promoting the transition to sustainability]. Ecosistemas, 16(1): 13–28. González de Molina, M., Petersen, P.F., Garrido Peña, F., & Caporal, F.R. (2020), Political agroecology: advancing the transition to sustainable food systems. Boca Raton, FL: CRC Press. González de Molina, M., Soto Fernández, D., Guzmán Casado, G., Infante-Amate, J., Aguilera Fernández, E., Vila Traver, J., & García-Ruiz, R. (2019), The social metabolism of Spanish agriculture, 1900-2008: the

Agroecology  13 Mediterranean way towards industrialization. Cham: Springer. Guzmán, G., & González de Molina, M. (2017), Energy in agroecosystems: a tool for assessing sustainability. Boca Raton, FL: CRC Press. Guzmán, G., González de Molina, M., & Sevilla, E. (eds.) (2000), Introducción a la agroecología como desarrollo rural sostenible [Introduction to agroecology as sustainable rural development]. Madrid: Ediciones Mundi-Prensa. Guzmán, G., López, D., Román, L., & Alonso, A.M. (2013), Participatory action research in agroecology: building local organic food networks in Spain. Agroecology and Sustainable Food Systems, 37(1): 127–46. https://​doi​.org/​10​ .1080/​10440046​.2012​.718997 Hecht, S.B. (1995), The evolution of agroecological thought. In M.A. Altieri (ed.), Agroecology: the science of sustainable agriculture. Boulder, CO: Westview Press, pp. 1–19. Klages, K.H.W. (1928), Crop ecology and ecological crop geography in the agronomic curriculum. Journal of the American Society of Agronomy 10: 336–53. Méndez, E., Bacon, C.M, & Cohen, R. (2013), Agroecology as a transdisciplinary, participa-

tory, and action-oriented approach. Agroecology and Sustainable Food Systems, 37(1): 3–18. Ploeg, J.D.V.D., Barjolle, D., Bruil, J., Brunori, G., Madureira, L.M.C., Dessein, J., Drąg, Z., Fink-Kessler, A., Gasselin, P., De Molina, M.G., et al. (2019), The economic potential of agroecology: empirical evidence from Europe. Journal of Rural Studies, 71: 46–61. https://​doi​ .org/​10​.1016/​j​.jrurstud​.2019​.09​.003 Toledo, V. (1999), Las ‘disciplinas híbridas’: 18 enfoques interdisciplinarios sobre naturaleza y sociedad [The ‘hybrid disciplines’: 18 interdisciplinary approaches to nature and society]. Persona y Sociedad, 13, 1. Toledo, V., & Barrera-Bassols, N. (2008), La memoria biocultural: la importancia ecológica de las sabidurías tradicionales [Biocultural memory: the ecological importance of traditional wisdom]. Barcelona: Icaria. Vandermeer, J. (2009), The ecology of agroecosystems. Sudbury, MA: Jones & Bartlett Learning. Wezel, A., Bellon, S., Doré, T., Francis, C., Vallod, D., & David, C. (2009), Agroecology as a science, a movement, and a practice. A review. Agronomy for Sustainable Development, 29(4): 503–15.

Manuel González de Molina

3. Agrowth 3.1

(GDP): that is, adopting an agrowth position (van den Bergh, 2011). An additional reason to support this position is that the measure of economic growth, GDP, is not a good proxy of social welfare or human progress. In this chapter I recast the main arguments for, and benefits of, an agrowth position, and discuss how it relates to other ideas about growth-vs-environment as developed under the umbrella of ecological economics.

Beyond the growth-vs-environment debate

Societies around the world have been unable to quickly and effectively respond to the vast inequity and sustainability challenges faced by humanity. This is partly due to voters and politicians suspecting that far-reaching social and environmental policies will reduce economic growth. What matters here is not whether there are social, biological, or physical limits to growth – we can endlessly discuss pro and contra arguments.1 More relevant than the facts are the beliefs and fears of people about growth impacts of ambitious policies. The modern growth debate is often considered to start with John Kenneth Galbraith (1958). He warned that American growth in the post-war period was too much driven by status and private gains, lacking investment in the public sector and policies to combat inequality. Of course, the classic economists were already concerned about limits to growth (Zweig, 1979), most notably in the person of Thomas Malthus. The catalysing publication for the growth debate’s focus since the 1970s on environmental and resource constraints is the “Limits to Growth” report to the Club of Rome by Meadows and colleagues (1972). Recently, the debate was refuelled through widespread attention for nine key planetary boundaries (Rockström et al., 2009), and particularly the challenges posed by climate change and policy (Antal and van den Bergh, 2016). The debate about environment versus growth is surrounded by uncertainties, notably as past facts do not predict the future, in turn explaining why distinct positions remain alive (Mishan, 1977; van den Bergh and de Mooij, 1999). While the past shows mainly environmentally damaging growth, it is unclear how far environmental policies might go in decoupling income from environmental pressure in the future. Rather than trying to settle this debate, it makes sense to adapt one’s views and strategies to the prevailing uncertainties. I have argued that a scientifically and politically logical approach is to be indifferent about gross domestic product

3.2

Motivations for agrowth

3.2.1 Doubts about pro- and anti-growth In the many discussions I have had over time about growth versus environment, I always have been critical of both strong pro- and anti-growth beliefs. Often this led others to put me in the opposite corner: that is, pro-growth painted me as anti-growth and vice versa. When I explained that I felt most comfortable with a “third position”, my co-debaters sometimes understood that, but often would quickly forget about it and return to the simple pro/contra-growth dichotomy. This motivated me to come up with a name for the third position, and at the time “agrowth” seemed most obvious to me – with the letter “a” before “growth” denoting being “agnostic” in the sense of “indifferent or neutral about growth”. To clarify my views, I will first explain why I think the polar opposites of green growth and degrowth are risky strategies. Optimism about green growth is understandable, as its implicit promise of “win–win” is hard to resist. But it requires an absolute decoupling of (overall) environmental pressure from aggregate income. This in turn involves a shift in investment from labour to resource productivity improvement, which will slow down labour productivity growth, in turn negatively affecting economic growth. Many people and politicians profess a belief in green growth, but often they merely pay lip service to it as they hesitate to support ambitious environmental policies out of fear these will harm growth. The opposite strategy is anti-growth in nature, regarding GDP growth as unconditionally undesirable. It proposes to foster a “downscaling of the economy” to meet environmental goals, nowadays often referred to as “degrowth” (Martínez-Alier et al., 2010; Schneider et al., 2010; Mastini et al., 2021). It is driven by an absolute belief that environmental pressures and GDP cannot be sufficiently decoupled, 14

Agrowth  15

or that improving resource efficiency and substitution in production and consumption will be inadequate to reduce environmental pressure. But this excludes the possibility that the economy adapts and restructures much in the face of strong policies, and that growth in the future will look quite differently from that in the past. It also raises the question which particular degrowth target or rate should be aimed for – difficult to answer, however, if the precise consequences of efficiency improvements and substitution are uncertain. Furthermore, degrowth thinking tends to confuse cause and effect:2 perhaps serious climate policy results in negative growth, temporarily or even permanently; but this does not mean the reverse is true: ex ante planning some degrowth pattern may not work out the best for well-being and climate. Last but not least, a degrowth strategy is problematic in terms of political feasibility. The term “degrowth” evokes negative associations in people’s minds (Raworth, 2015; Drews and Antal, 2016; Drews and Reese, 2018) by creating a mixture of anti-growth and anti-capitalism messages which suggest that major sacrifices are needed to protect the environment. In view of human psychology, this could well undercut, rather than build, political support for effective environmental policies. 3.2.2 “Beyond GDP” implies agrowth It is widely accepted in economics, other social sciences, and environmental sciences that GDP cannot serve as an indicator of societal progress under all circumstances and in all development phases. Three important reasons are: (1) it does not capture income inequality, relative income effects, and adaptation to higher incomes; (2) it neglects the informal economy; and (3) it excludes environmental externalities, ecosystem damage, and depletion of natural resources. However, this recognition has not translated into the demise of GDP in society and policy. One reason is that the majority of economists, journalists, and politicians, irrespective of political affiliation, still express themselves rather uncritically when it comes to GDP. Another reason is that, while many alternative measures have been proposed, none of these has been embraced by a sufficiently wide group of practitioners (Costanza et al., 2014).

As a result, GDP remains dominant in macroeconomics and public policy debates. Its shortcomings as a welfare measure mean that using GDP as a progress indicator causes an “information failure”, which can steer the economy in a very wrong direction in both social-welfare and environmental terms –under both absolute pro- and anti-growth strategies.3 Most politicians, nevertheless, routinely use GDP as if it were a good proxy of welfare. This “GDP paradox” meaning a bad progress indicator that continues to have much influence on politics and policy (van den Bergh, 2009), warrants fundamental changes in our thinking. An agrowth position is one way to accomplish this, being a logical consequence of accepting the shortcomings of the GDP indicator, and emitting the clear message to ignore it.

3.3 Characterizing agrowth in distinct ways

By not choosing between “anti-growth” and “pro-growth”, agrowth adds a third neutral position that is indifferent about GDP growth. This comes down to removing the constraint of “always or unconditional economic growth” – which others have referred to as “growthmania” (Mishan, 1967) and “growth fetishism” (Stiglitz, 2009). One constraint less implies more flexibility for politicians to address important social and environmental problems, as they no longer have to assume that solutions go hand in hand with GDP increases. Clear examples of urgent problems are poverty and inequality, structural unemployment, and global environmental change. Since unconditional GDP growth often acts as a constraint on solving these, relieving the continuous-growth constraint facilitates the implementation of genuine solutions. It is important to realize that an agrowth position means one can be concerned or critical about growth without this resulting in a radical and dogmatic anti-growth position. Under an agrowth strategy, it is possible that periods with high growth are followed by low or even negative growth, provided they go along with progress in welfare terms. Under an agrowth paradigm, we would, if necessary – without being aware due to ignoring GDP information – sacrifice some GDP growth in exchange for a better environment, more equality, or more leisure. So agrowth means Jeroen van den Bergh

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prioritizing welfare over income and getting rid of the idea that growth (or degrowth) is required or sufficient for societal progress. Indeed, agrowth is consistent with the more subtle view that, to stay within environmental limits, we need selective growth and selective decline triggered by environmental policies,

agrowth strategy does not, ex ante, exclude any of these areas. The case of Figure 3.1a (conflict) illustrates that a green-growth strategy aiming for growth beyond the rate g is not wise as it will not achieve its aim of ending up in area C. The reason is illustrated by the point

Note: Search space for human progress spanned by relative changes in GDP and environmental sustainability (ES) in interval [t, t+1]; bold letters denote the rectangles separated by the vertical and horizontal broken lines. Source: Figures 2 and 3 in van den Bergh (2018).

Figure 3.1

Green growth, degrowth, and agrowth under growth–environment conflict (a) and compatibility (b)

and not indiscriminate growth or degrowth. In addition, agrowth will help to avoid disappointment in rich countries about low and likely declining growth rates in the future due to declining returns on investments in education and technological innovation (Gordon, 2012). Agrowth is a strategy that is more precautionary and robust in avoiding regret under uncertain futures than are pro- and anti-growth absolutism. This is illustrated by Figure 3.1, panels a and b, with the first depicting conflict between growth and social or environmental goals (indicated by feasible outcomes below curve 1) and the second showing compatibility (feasible outcomes below curve 2). The graphs plot the GDP growth rate (horizontal axis) against the change in other components of human welfare, including environmental sustainability (ES; vertical axis). Now, a degrowth strategy strives to be in (rectangular) area A, a low-growth strategy in area B, and a high-growth strategy in area C (where growth is higher than rate g, such as the popular aim of 2 per cent growth). As opposed to the previous strategies, an Jeroen van den Bergh

above the constraint 1, which represents an infeasible goal. If one strives for high growth associated with it, the economy will end up in the point below the constraint (following the arrow down). In this case, degrowth (area A) and low-growth (area B) strategies are feasible. But as Figure 3.1b (compatibility) illustrates, a degrowth strategy can also lead to decreases in human welfare (e.g. when tax revenues are insufficient to finance important public goods). Indeed, trying to be in area A fails here, as one will be forced to be below constraint 2, indicated by an arrow from the infeasible point above the curve to the feasible point below it (again following the arrow down). In this case, a high-growth strategy is feasible. Taking Figures 3.1a and b jointly, only an agrowth strategy is robust and feasible under any (uncertain) circumstance – growth/environment conflict or compatibility – as it does not limit its search to one of areas A, B, and C. For more discussion, see van den Bergh (2017, 2018). Two studies provide empirical evidence suggesting that an agrowth viewpoint can already count on considerable support. One

Agrowth  17

involved a survey with a sample representative of the general population of Spain (N = 1008), and another undertook a survey among researchers (N = 814) from economics, environmental social sciences, and natural sciences (Drews and van den Bergh, 2016, 2017). Figure 3.2 shows aggregate results for both, indicating that agrowth is more popular already than degrowth, and among scientists not so far behind green growth even.

Source:

Figure 3.3 in van den Bergh and Drews (2019).

Figure 3.2

Scientists’ vs citizens’ preferences for a growth-vs-environment strategy

While green growth has always been popular, and degrowth has quickly gained interest in a short time, agrowth seems to follow a more gradual path of rising support. In a recent study we show that, since 2014, the share of people supporting green growth has fallen, while the share of agrowth has risen (Savin et al., 2021). In addition, many people from distinct corners and disciplines, including people whom I respect much, have expressed in personal communication with me that they appreciate or even can embrace the idea of agrowth, when previously they leaned towards the pro- or anti-growth position. I take this as a further indication that agrowth may contribute to depolarizing the growth debate and arrive at an agreement about which strategy makes the most sense scientifically and politically in achieving ambitious social and sustainability goals.

3.4

continued growth in the face of limited environmental and resource capacities to sustain pollution, waste, and resource extraction. The most consistent scholar in this respect is, without any doubt, Herman Daly. Through his precise ideas and attractive writing, he influenced the thinking of various generations of economists and environmental scientists, including me. In the 1960s and ’70s, he developed the notion of a steady-state

Relation with other ideas in ecological economics

An important motivation for the emergence of ecological economics was a concern about

economy aimed at keeping the throughput of energy and materials within safe environmental limits (Daly, 1968, 1977). Later he elaborated the related idea of an optimal physical scale of the economy using the analogy of a plimsoll line in a boat to avoid sinking by overweight. Hence, he suggested that the notion of optimal scale needs serious attention from economists next to that of optimal allocation (Daly, 1992). In line with these ideas, Daly has been critical of the GDP indicator, proposing an alternative measure, namely, the Index of Sustainable Economic Welfare (ISEW; Daly and Cobb, 1989), which has seen much application (Posner and Costanza, 2011). One may wonder if Daly’s views on growth are consistent with an agrowth position. In my interpretation they are at least not inconsistent. Daly’s steady-state economy is focused on the physical dimensions of the economy, which he clearly distinguishes from the monetary dimension, without assuming a one-to-one connection. Daly believes that decoupling may be very difficult, but also Jeroen van den Bergh

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thinks the economy has much flexibility in responding to environmental regulation, and hence does not exclude the possibility that GDP might increase when material- and energy-extensive activities become more dominant. Kerschner (2010) further argues that Daly’s steady-state economy, with its “air of top-down decision-making”, is complementary to degrowth, with its “focus on grass-roots initiatives”. In addition, Daly’s work on an alternative measure (ISEW) to replace GDP matches well with an agrowth perspective stressing to ignore GDP information. Another intellectual father of ecological economics, Nicholas Georgescu-Roegen (1977), was critical of Daly’s steady-state economy. His judgement followed from his view that the economy is a thermodynamic and evolutionary, rather than a mechanical, system. His specific thesis was that 100 per cent recycling of materials is impossible, implying that the economic system will ultimately run out of low-entropy materials as these become fully dissipated and unavailable. In line with this, Georgescu-Roegen has been baptized “the father of degrowth” (Martínez-Alier et al., 2010). The validity and relevance of his thesis have, though, been disputed from various theoretical and empirical angles (Ayres, 1997, 1999; Cleveland and Ruth, 1997; van den Bergh, 2020). According to Ayres, the best we can say is that not all materials in the Earth can be simultaneously in active service at any given time, because even the most efficient recycling generates wastes. This does, however, not mean that the economy will run out of useful materials. Since the Earth is an energetically open system, solar energy can fuel continuous recycling at a positive percentage (even though less than 100 per cent) of all materials in the “wastebasket”. The latter, then, actually becomes an ore. Of course, slow technological progress and economic rationale going for the new instead of reuse and repair may result in deviating from the physically most sustainable path. Here many uncertainties enter into the picture, underpinning a choice for agrowth.

anti-growth has managed to create sufficient political support for implementing necessary sustainability policies. Agrowth may help to turn the tide. An important advantage of agrowth is that, through reducing polarization, it can weaken political resistance against, or even create agreement about, serious environmental and climate policies. Unlike unconditional pro-growth, agrowth does not give priority to income growth over environment or equity. The option of agrowth may change the nature of the growth debate from searching for truth to adopting precaution in the face of a combination of planetary boundaries, a complex economy, and uncertainty about the future. Degrowth is sometimes confused with agrowth – perhaps as they both are new concepts that emerged from being critical of growth. But whereas degrowth is moral or normative and multi-interpretable (van den Bergh, 2011; Jakob and Edenhofer, 2014; Eversberg and Schmelzer, 2018), agrowth is logical and specific, as it is directly linked to the GDP indicator: it does not serve as a good progress measure, so ignore its dynamics – which implies indifference about GDP growth and decline. Another term used is “post-growth”. This is, however, problematic from a scientific perspective as distinct authors interpret it very differently (Jackson, 2021; Petschow et al., 2020) or as agrowth (Lehmann et al., 2022). What distinguishes agrowth from degrowth and other post-growth notions is that one can be critical of GDP growth without being against it. Several actions can foster a transition to an agrowth paradigm (van den Bergh, 2017): interrogate the preoccupation with growth in journalism, policy circles, and politics; give attention to the shortcomings of GDP as a welfare/progress indicator in all education; and encourage influential global players, such as the International Monetary Fund, the Organisation for Economic Co-operation and Development, and the World Bank, to embrace an agrowth position. More generally, I invite everyone to reconsider their pro- or anti-growth views after recognizing that there is a third option.

3.5 Conclusions

Acknowledgements

I have argued that an agrowth position is logical from both scientific and political perspectives. Neither green growth nor Jeroen van den Bergh

I am grateful to Stefan Drews and Joël Foramitti for feedback, and to Joan Martínez-

Agrowth  19

Alier for discussions on the topic of Section 3.4. Jeroen van den Bergh

Notes 1.

2.

3.

One should distinguish limits to the economy from limits to growth or scale. The first do not necessarily imply the second, as substitution of inputs, more efficient technologies and practices, and changes in the composition of consumption are additional ways to stay within the economy’s physical and biological limits. In fact, most green-growth believers accept limits and planetary boundaries, or admit that we are currently exceeding these, but are optimistic that the mentioned mechanisms can keep our economy within the limits. In line with this confusion, it is not uncommon to see uses of the term “degrowth” to denote negative GDP growth as an outcome, when “economic decline” and “recession” are more precise and common terms which help to clearly distinguish an unintended ex post outcome (e.g. of social or environmental policies) from an ex ante degrowth strategy. In support of this assessment, Jakob and Edenhofer (2014) state that both pro- and degrowth “constitute inadequate foundations for public policy as they fail to appropriately conceptualize social welfare” (p. 447).

References

Antal, M., and J.C.J.M. van den Bergh (2016). Green growth and climate change: Conceptual and empirical considerations. Climate Policy 16(2): 165–77. Ayres, R.U. (1997). Comments on Georgescu-Roegen. Ecological Economics 22(3): 285–87. Ayres, R.U. (1999). The second law, the fourth law, recycling and limits to growth. Ecological Economics 29(3): 473–83. Cleveland, C.J., and M. Ruth (1997). When, where, and by how much do biophysical limits constrain the economic process? A survey of Nicholas Georgescu-Roegen’s contribution to ecological economics. Ecological Economics 22: 203–23. Costanza, R., I. Kubiszewski, E. Giovannini, H. Lovins, J. McGlade, K.E. Pickett, K.V. Ragnarsdóttir, D. Roberts, R. De Vogli, and R. Wilkinson (2014). Time to leave GDP behind. Nature 505: 283–5. Daly, H.E. (1968). On economics as a life science. Journal of Political Economy 76: 392–406. Daly, H.E. (1977). Steady-State Economics: The Political Economy of Biophysical Equilibrium

and Moral Growth. W.H. Freeman, San Francisco. Daly, H.E. (1992). Allocation, distribution, and scale: Towards an economics that is efficient, just, and sustainable. Ecological Economics 6: 185–93. Daly, H.E., and W. Cobb (1989). For the Common Good: Redirecting the Economy Toward Community, the Environment and a Sustainable Future. Beacon Press, Boston. Drews, S., and M. Antal (2016). Degrowth: A “missile word” that backfires? Ecological Economics 126: 182–7. Drews, S., and G. Reese (2018). “Degrowth” vs. other types of growth: Labeling affects emotions but not attitudes. Environmental Communication 12: 763–72. Drews, S., and J. van den Bergh (2016). Public views on economic growth, the environment and prosperity. Global Environmental Change 39: 1–14. Drews, S., and J. van den Bergh (2017). Scientists’ views on economic growth versus the environment. Global Environmental Change 46: 88–103. Eversberg, D., and M. Schmelzer (2018). The degrowth spectrum. Environmental Values 27: 245–67. Galbraith, J.K. (1958). The Affluent Society. Houghton Mifflin Company, Boston. Georgescu-Roegen, N. (1977). The steady-state and ecological salvation. Bioscience 27: 266–70. Gordon, R.J. (2012). Is U.S. economic growth over? Vox, 11 September. https://​ voxeu​ .org/​ article/​us​-economic​-growth​-over Jackson, T. (2021). Post Growth—Life After Capitalism. Polity Press, Cambridge, UK, and Medford, MA. Jakob, M., and O. Edenhofer (2014). Green growth, degrowth, and the commons. Oxford Review of Economic Policy 30(3): 447–68. Kerschner, C. (2010). Economic de-growth vs. steady-state economy. Journal of Cleaner Production 18: 544–51. Lehmann, C., O. Delbard, and S. Lange (2022). Green growth, a-growth or degrowth? Investigating the attitudes of environmental protection specialists at the German Environment Agency. Journal of Cleaner Production 336: 130306. Martínez-Alier, J., U. Pascual, F.-D. Vivien, and E. Zaccai (2010) Sustainable de-growth: Mapping the context, criticisms and future prospects of

Jeroen van den Bergh

20  Elgar encyclopedia of ecological economics an emergent paradigm. Ecological Economics 69(9): 1741–7. Mastini, R., G. Kallis, and J. Hickel (2021). A green new deal without growth? Ecological Economics 179: 106832. Meadows, D.H., D.L. Meadows, J. Randers, and W.W. Behrens III (1972). The Limits to Growth. Universe Books, New York. Mishan, E.J. (1967). The Costs of Economic Growth. Praeger, New York. Mishan, E.J. (1977). The Economic Growth Debate: An Assessment. George Allen & Unwin, London. Petschow, U., S. Lange, D. Hofmann, E. Passarkoi, N. aus dem Moore, T. Korfhage, A. Schoofs, and H. Ott (2020). Social well-being within planetary boundaries: The precautionary post-growth approach. Report 234/2020, German Environmental Agency, Berlin. Posner S.M., and R. Costanza (2011). A summary of ISEW and GPI studies at multiple scales and new estimates for Baltimore City, Baltimore County, and the State of Maryland. Ecological Economics 70(11): 1972–80. Raworth, K. (2015). Why degrowth has out-grown its own name. Guest post by Kate Raworth. Oxfam Blogs. https://​frompoverty​.oxfam​.org​ .uk/​why​-degrowth​-has​-out​-grown​-its​-own​ -name​-guest​-post​-by​-kate​-raworth/​ Rockström, J., W. Steffen, K. Noone, Å. Persson, F.S. Chapin III, E. Lambin, T.M. Lenton, M. Scheffer, C. Folke, H. Schellnhuber, B. Nykvist, C.A. De Wit, T. Hughes, S. van der Leeuw, H. Rodhe, S. Sörlin, P.K. Snyder, R. Costanza, U. Svedin, M. Falkenmark, L. Karlberg, R.W. Corell, V.J. Fabry, J. Hansen, B. Walker, D. Liverman, K. Richardson, P. Crutzen, and J. Foley (2009). Planetary boundaries: Exploring the safe operating space for humanity. Ecology and Society 14(2): 32. http://​ www​.ecologyandsociety​.org/​vol14/​iss2/​art32/​ Schneider, F., G. Kallis, and J. Martínez-Alier (2010). Crisis or opportunity? Economic

Jeroen van den Bergh

degrowth for social equity and ecological sustainability. Journal of Cleaner Production 18(6): 511–18. Stiglitz, J.E. (2009). GDP fetishism. The Economists’ Voice 6(8): 5. http://​www​.bepress​ .com/​ev/​vol6/​iss8/​art5 Savin, I., S. Drews, and J. van den Bergh (2021). GEM: A short “growth-vs-environment” module for survey research. Ecological Economics 187: 107092. van den Bergh, J. (2009). The GDP paradox. Journal of Economic Psychology 30(2): 117–35. van den Bergh, J. (2011). Environment versus growth – A criticism of “degrowth” and a plea for “agrowth”. Ecological Economics 70(5): 881–90. van den Bergh, J. (2017). A third option for climate policy within potential limits to growth. Nature Climate Change 7(February): 107–12. van den Bergh, J. (2018). Agrowth instead of anti- and pro-growth: Less polarization, more support for sustainability/climate policies. Journal of Population and Sustainability 3(1): 53–74. van den Bergh, J. (2020). Six policy perspectives on the future of a semi-circular economy. Resources, Conservation & Recycling 160: 104898. van den Bergh, J., and S. Drews (2019). Green “agrowth” – The next development stage of rich countries, in Handbook on Green Growth, edited by R. Fouquet. Edward Elgar Publishing, Cheltenham, UK, and Northampton, MA, pp. 52–66. van den Bergh, J., and R.A. de Mooij (1999). An assessment of the growth debate, in Handbook of Environmental and Resource Economics, edited by J. van den Bergh. Edward Elgar Publishing, Cheltenham, UK, and Northampton, MA, pp. 643–55. Zweig, K. (1979). Smith, Malthus, Ricardo, and Mill: The forerunners of limits to growth. Futures 11(6): 510–23.

4. Anthropocene

While evidence of the various “geological signals” of the Anthropocene have become well established, the official naming of a new epoch ultimately needs super-majority (> 60 percent) support of the AWG and its parent scientific bodies, as well as ratification by the Executive Committee of the International Union of Geological Sciences. A proposal under development concludes that dating the Anthropocene would be “optimally placed in the mid-20th century, coinciding with the array of geological proxy signals preserved within recently accumulated strata and resulting from the ‘Great Acceleration’ of population growth, industrialization and globalization” (Subcommission on Quaternary Stratigraphy, n.d.). Specifically, the AWG is recommending that the “sharpest and most globally synchronous of these signals, that may form a primary marker, is made by the artificial radionuclides spread worldwide by the thermonuclear bomb tests from the early 1950s” (see Figure 4.1). The post-1950 Great Acceleration associated with the Anthropocene is characterized by key socioeconomic and Earth systems trends documented by the IGBP. Table 4.1 compares the magnitude of change over six decades of accelerating socioeconomic and Earth system trends with the two preceding 60-year time frames. The only trends that have slowed since 1950 are the number of active large dams, land area in agriculture, plus associated fertilizer consumption. The other “hockey stick” diagrams tracked by the IGBP through 2010 include a global temperature increase of 0.47oC (since a 1943 base year), Antarctic stratospheric ozone loss around 60 percent (seemingly stabilizing since the 1990s), and an accelerating decrease in terrestrial species abundance (from a 1950 mean decrease of 14 percent to a 29 percent decrease in 2000). The loss of species during the Anthropocene is also associated with a sixth mass extinction of the Phanerozoic “visible life” Eon. Ceballos et al. (2015) estimate a conservative 20th-century extinction rate for vertebrates of up to 100 times greater than a “background” rate. In the current century, a 2019 global assessment by the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services reviewed over 15 000 scientific publications and a large body of Indigenous and local knowledge, concluding that “around 1 million species already face

Earth system scientists have proposed the Anthropocene epoch – the age of humans – as a new geochronological time interval within the Quaternary Period of the Cenozoic Era. Popularized by Nobel-prize-winning chemist Paul Crutzen, this new time unit would be stratigraphically distinct from the Holocene, a post-glacial epoch that began 10 000–12 000 years ago. Crutzen and Stoermer (2000) first argued for naming the Anthropocene in a newsletter of the International Geosphere– Biosphere Programme (IGBP) “to emphasize the central role of mankind in geology and ecology.” The consideration of this new epoch – the second smallest geochronological unit used by geologists in dating rock layers – has been taken up by various scientific bodies, most notably an “Anthropocene Working Group” (AWG) of the International Commission of Stratigraphy formed in 2009. The AWG recognizes many “phenomena” associated with global human impact, many of which are “altering the trajectory of the Earth System” and “are being reflected in a distinctive body of geological strata” (Subcommission on Quaternary Stratigraphy, n.d.). These include: ● An order-of-magnitude increase in erosion and sediment transport associated with urbanization and agriculture; ● Marked and abrupt anthropogenic perturbations of the cycles of elements, such as carbon, nitrogen, phosphorus, and various metals together with new chemical compounds; ● Environmental changes generated by these perturbations, including global warming, sea-level rise, ocean acidification, and spreading oceanic “dead zones”; ● Rapid changes in the biosphere, both on land and in the sea, as a result of habitat loss, predation, explosion of domestic animal populations, and species invasions; and ● The proliferation and global dispersion of many new “minerals” and “rocks,” including concrete, fly ash and plastics, and the myriad “technofossils” produced from these and other materials (Subcommission on Quaternary Stratigraphy, n.d.).

21

22  Elgar encyclopedia of ecological economics

Source:

Adapted from Head et al. (2021).

Figure 4.1

The proposed Anthropocene epoch

extinction, many within decades” (p. xvi). That’s nearly one in nine species slated to vanish from the Earth in a blink of geological time. The causes and consequences of this human-dominated epoch go well beyond the empirics of global environmental impact and geochronological dating. The Anthropocene discourse has also been explicitly normative. For example, Crutzen and Schwägerl (2011) have proposed a global call to action, where: We must change the way we perceive ourselves and our role in the world. Students in school are still taught that we are living in the Holocene, an era that began roughly 12 000 years ago at the end of the last Ice Age. But teaching students that we are living in the Anthropocene, the Age of Men, could be of great help. Rather than representing yet another sign of human hubris, this name change would stress the enormity of humanity’s responsibility as stewards of the Earth.

The resulting discourse over “humanity’s responsibility” in the Anthropocene Jon D. Erickson

has been characterized between poles of hope and despair. For example, Bennett et al.’s (2016) interpretation of a “good” Anthropocene stresses “new strategies for creating a more just, prosperous, and ecologically diverse world,” whereas dystopian views emphasize “irreversible environmental degradation and societal collapse that ultimately diminish human quality of life” (p. 441). A critical discourse has also opened around the human-centricity of naming the Anthropocene, especially over ecomodernist and theocratic views of humanity’s “ability to transform and control nature” (Hamilton, 2016, p. 233). Other critiques have stressed the need to exit the Anthropocene and enable a new era, for instance, an “Ecozoic” proposed by Thomas Berry where humans live in a mutually enhancing relationship with the community of life (Vargas Roncancio et al., 2019). There are also complimentary discourses that stress the cultural aspects of a human-dominated epoch. For example, ecological economist Richard Norgaard (2015)

Anthropocene  23 Table 4.1

The Great Acceleration of the Anthropocene epoch

Global Trend

Base

Total Percent Change

Units

1830–1890

1890–1950

1950–2010

Socioeconomic System

 

 

 

 

Population

#

35%

64%

173%

  Real GDP

USD

454%

294%

806%

  Foreign Direct Investment

USD

n/a

n/a

636 091%

  Urban Population

#

106%

241%

376%

  Primary Energy Usea

exajoule

71%

166%

375%

  Fertilizer Consumptionb

tons

n/a

1 608%

1 081%

  Large Dams

# > 15 meters

408%

854%

449%

  Water Usec

kilometer3

n/a

83%

216%

  Paper Productiond

tons

n/a

n/a

438%

 Transportatione

vehicles

n/a

n/a

623%

 Telecommunications

lines/subscriptions

n/a

n/a

909 725%

  International Tourism

arrivals

n/a

n/a

3 615%

Earth System

  

  

  

  

  Carbon Dioxide

parts per million

3.2%

6.2%

24.1%

  Nitrous Oxide

parts per million

2.3%

3.8%

12.2%

 Methane

parts per billion

11.5%

31.1%

56.7%

  Ocean Acidificationf

H+ concentration

2.3%

4.4%

17.8%

  Marine Fish

tons

n/a

n/a

355%

  Shrimp Aquaculture

tons

n/a

n/a

285 855%

  Tropical Forest Loss

km2

−4.0%

−10.0%

−13.6%

  Agricultural Land Gain

km2

54.2%

65.1%

21.3%

Source: IGBP (2014). Notes: n/a = data not available. a Third period is 58 years (1950–2008). Data for 1830 are extrapolated from 1800 and 1850 estimates. b Second period is 50 years (1900–50). Data are not reported before 1900. c Second period is 49 years (1901–50). Data are not report before 1901. d Third period is 49 years (1961–2010). Data are not reported before 1961. e Third period is 47 years (1963–2010). Data are not reported before 1963. f First period is 40 years (1850–90). Data are not reported before 1850.

has suggested the “Econocene” as a fitting description of our time, emphasizing “the expanding market economy, the ideological system that supports it, and its impact on society and the environment” (p.  1). This narrative is rooted in “economism,” what Norgaard defines as “the reduction of all social relations to market logic” (p. 1). Similar arguments have been made for the Capitalocene (Moore, 2016) and Technocene (Martins, 2018). For ecological economics, the Anthropocene discourse aligns with long-standing arguments to study the economy in relationship with the biophysical limits of the Earth system. For example, Daly’s (1990) call for a “full-world economics” emphasized a new era in which “natural capital will be increasingly limited” (p. 4).

Natural capital is necessary for both the accelerating human use of energy and materials in the Anthropocene, as well as the services provided from intact ecosystems, such as waste assimilation, habitat, and climate stabilization. Ecological economics stresses the complementarity of natural capital to manufactured and human capital in economic production, with biophysically determined limits on substitutability. Where there is less clarity from ecological economics is on how to respond to the divergent cultural narratives of the Anthropocene. Brown and Erickson (2016) specifically call for an “economics for the Anthropocene” that would “frame economic production as biophysical transformation towards socially constructed ends” (p. 46). While ecological economists generally agree on the biophysical Jon D. Erickson

24  Elgar encyclopedia of ecological economics

framing of the Anthropocene discourse, there is less agreement on its normative implications. Wironen and Erickson (2020) explore the “unresolved tension regarding modern and postmodern social theory” within the transdisciplinary field, and suggest “a critically modern ecological economics could draw on aligned social movements and build on deliberative theory as a foundation for social and political change fit for navigating the Anthropocene” (p. 62). Ultimately, beyond the science of documenting a new epoch, it may be in deliberating the meaning of the Anthropocene across diverse thought, opinion, and experience where humanity may shape and legitimize a new social reality. Jon D. Erickson

References

Bennett, E.M., Solan, M., Biggs, R., MacPhearson, T., Norstrom, A., Olsson, P., Pereira, L., et al. “Bright Spots: Seeds of a Good Anthropocene,” Frontiers in Ecology and Environment 14(8): 441–8, 2016. https://​doi​.org/​10​.1002/​fee​.1309 Brown, P.G., and Erickson, J.D. “How Higher Education Imperils the Future: An Urgent Call for Action,” Balance 2: 42–8, 2016. Ceballos, G., Ehrlich, P.R., Barnosky, A.D., García, A., Pringle, R.M., and Palmer, T.M. “Accelerated Modern Human-Induced Species Losses: Entering the Sixth Mass Extinction,” Science Advances 1(5): e1400253, 2015. Crutzen, P.J., and Stoermer, E.F. “The Anthropocene,” International Geosphere– Biosphere Programme Newsletter 41, 2000. Crutzen, P.J., and Schwägerl, C. “Living in the Anthropocene: Toward a New Global Ethos,” Yale Environment 360, January 24, 2011. Daly, H., “Toward Some Operational Principles of Sustainable Development,” Ecological Economics 2: 1–6, 1990. Hamilton, C. “The Theodicy of the ‘Good Anthropocene,’” Environmental Humanities 7(1): 233–8, 2016. Head, M.J., Steffen, W., Fagerlind, D., Waters, C.N., Poirier, C., Syvitski, J., Zalasiewicz,

Jon D. Erickson

J.A., Barnosky, A.D., Cearreta, A., Jeandel, C., Leinfelder, R., McNeill, J.R., Rose, N.L., Summerhayes, C., Wagreich, M., and Zinke, J. “The Great Acceleration is Real and Provides a Quantitative Basis for the Proposed Anthropocene Series/Epoch,” Episodes: 021031, 2021. International Geosphere-Biosphere Programme. “The Great Acceleration,” Excel spreadsheet, version 2014. Accessed on August 3, 2022 at: http://​www​.igbp​.net/​globalchange/​ greatacceleration​.4​.1b8ae20​512db692f2​ a680001630​.html Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services. Global Assessment Report on Biodiversity and Ecosystem Services of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services, IPBES Secretariat, Bonn, Germany, 2019. Martins, H. The Technocene: Reflections on Bodies, Minds, and Markets, Anthem Press, New York, 2018. Moore, J.W. (Ed.), Anthropocene or Capitalocene? Nature, History, and the Crisis of Capitalism, PM Press, Oakland, CA, 2016. Norgaard, R. “The Church of Economism and Its Discontents,” Great Transition Annual Review, 2015(2), 2015. https://​ greattransition​ .org/​publication/​the​-church​-of​-economism​-and​ -its​-discontents Subcommission on Quaternary Stratigraphy. “Working group on the ‘Anthropocene,’” n.d. http://​quaternary​.stratigraphy​.org/​working​ -groups/​anthropocene/​ Vargas Roncancio, I., Temper, L. Sterlin, J. Smolyar, N.L., Sellers, S., Moore, M., Melgar-Melgar, R., Larson, J., Horner, C., Erickson, J.D., Egler, M., Brown, P.G., Boulot, E., Beigi, T., and Babcock, M. “From the Anthropocene to Mutual Thriving: An Agenda for Higher Education in the Ecozoic,” Sustainability 11(12): 3312, 2019. Wironen, M.B., and Erickson, J.D. “A Critically-Modern Ecological Economics for the Anthropocene,” The Anthropocene Review 7(1): 62–76, 2020.

5. Biodiversity conservation

scientific concept: it is perceived as a value, or as having a value. It is a widely recognised fact, regardless of which of the above groups one is in, that global biodiversity is in peril. For at least four decades, there has been growing concern about the rate at which species are being lost from ecosystems. During this time, ‘remarkable progress has been made towards understanding how the loss of biodiversity affects the functioning of ecosystems and thus affects society’ (Cardinale et al., 2012, 59), but the losses keep rising. It is no coincidence that interest in biodiversity loss has been at the top of the political agenda in recent decades. Although scientific evidence shows that the rate of species extinction has increased since the 19th century, during the last half-century it has skyrocketed (Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services [IPBES], 2019). This drastic increase has to do with the fact that humanity—with our ever-growing economic activity and social metabolism—is now the most dominant influence on nature (IPBES, 2019). Biodiversity conservation is an action we take to ease the negative consequences of economic activity on nature. The relentless pursuit of the world’s economies to achieve sustained economic growth—through a growing consumption of natural resources and increasing emissions—has led to this progressive loss of biodiversity. Yet biodiversity conservation, as we currently interpret and implement it, is dissociated from the economic system. Conservation emerges at the moment when the consequences— or externalities, as mainstream economists call them—are already evident and, in many cases, irremediable. This dissociation between the cycles and reproduction of the economic system and nature has led many experts to advocate for effective solutions through the incorporation of biodiversity into the market economic system. The logic of these proposals rests on the standpoint that the market is the best allocator of resources—including biodiversity, especially as we have commodified it. As has been widely argued (IPBES, 2019; Otero et al., 2020), the relationship is very simple: economic growth degrades biodiversity. As long as conservation ceases to be merely a palliative action, and the economic system is restricted within a sustainable scale based

5.1 Introduction

Biodiversity is a shortening of the term ‘biological diversity’. The simplest definition is that it is a synonym for the variety of life on Earth, the very fabric of life (Shiva, 2000). The most quoted definition is that of the United Nations Convention on Biological Diversity (CBD, 1992, art. 2): ‘“Biological diversity” means the variability among living organisms from all sources including, inter alia, terrestrial, marine and other aquatic ecosystems and the ecological complexes of which they are part; this includes diversity within species, between species and of ecosystems’. Before going into the specifics of how societies have cared for or addressed the loss of biodiversity, it is important to note that we do not all understand the same thing when we read or hear the term biodiversity. According to Gaston (1996), although its meaning is usually taken to be universal, biodiversity is interpreted differently by different groups of people. First, there are those who regard biodiversity as a concept. For them, expressed as the ‘variety of life’, biodiversity is essentially a wide-ranging abstract concept that can be distinguished at different levels of biological organisation: genetic, species, and ecosystems. Then, there are those who regard it as a measurable entity rather than simply an abstract concept (e.g. species richness, relative abundance, composition, presence/absence of key species). In choosing a particular aspect of biodiversity to measure, value is placed on that aspect for the purposes of the exercise at hand. Finally, there are those who regard it as predominantly a social or political construct, and there is general agreement that biodiversity, per se, is a good thing, that its loss is bad, and consequently, that something should be done to maintain it. These interpretations have been made to the extent that the term ‘biodiversity’ is seen to embody not only the variety of life, but also the importance of that variety, the crisis represented by its loss, and the need for conservation action. In this wider arena, biodiversity is not a neutral 25

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on the laws of matter and energy, then we will be able to generate schemes where economic activity does not increase the biodiversity loss we are currently experiencing.

5.2 Values and approaches

Since the formal establishment of the field of conservation—which happened gradually over the course of the 20th century— and particularly with the emergence of the sub-discipline of conservation biology in the 1980s, biodiversity conservation has been linked to the natural sciences. In the quest to save and protect the planet’s biodiversity, concerned biologists have triggered the establishment of what is often referred to as the ‘crisis discipline’ (Soulé, 1985). For decades, the narrative around biodiversity conservation was constructed by biologists who, given their interest in moving from knowledge generation to action, were placed in governmental conservation institutions or nongovernmental organisations (NGOs). These biologists and institutions were responsible for outlining the guidelines of what to conserve, where, and how. Whether in the trenches of academia, the public sector, or NGOs, conservation biologists have succeeded in defining conservation agendas and making decisions on how to implement them. Conservation scientists have kept the narrative of their work on two tracks: while motivated by the goal of protecting biodiversity, they are driven by curiosity, the joy of discovery, and value-free objectivity (Kareiva et al., 2018). However, the central concern of the activity has to do with the word that comes after the concept of biodiversity—conservation. Independently of what we understand by biodiversity, conservation itself is an action or a process. This action is carried out by human beings in a specific context: biophysical, geographical, historical, political, social, and cultural. Thus, as Alcorn (1994, 11) states, ‘[w]hile proof of conservation success is ultimately biological, conservation itself is a social and political process, not a biological process’. Following this interpretation, if we want to succeed in our conservation aims, we are obliged to change how decisions are made and policies designed—and by whom. Funtowicz and Ravetz (1991) have already argued that when facts are uncertain, values in dispute, stakes high, and decisions urgent, Eduardo García-Frapolli

we have to extend the decision-making process to a wider community. It is important to note, too, that the inclusion of other disciplines is not entirely unrelated to how conservation biology was conceptualised and how it has been put into practice. Over the decades, conservation biology has been breaking down disciplinary barriers, especially within the natural sciences. However, probably because of the limited success conservation biology has had in achieving its main objective, it has been necessary to incorporate other disciplines. Nevertheless, those who decide how we should conceptualise and implement conservation continue to be in the natural sciences, in fields that are dominated, to a large extent, by the ‘Western conservation movement’ (Pascual et al., 2021) and the scientisation of policy making. The Western conservation movement has been largely successful in defining the conservation agenda, for example, the approach of conserving rare species or wild ecosystems. Where it has been most successful is in the idea of conserving the wilderness or pristine landscapes through protected areas (PAs). PAs are widely believed to be one of the most effective means of reducing global biodiversity loss (Schulze et al., 2018). The strength of the discourse surrounding PAs, and their effectiveness in reducing biodiversity loss, is reflected in the fact that the international community recurrently commits to increasing the percentage of conserved land and sea areas globally through well-connected systems of PAs. In the latest international agreement (CBD, 2021), there is already mention of conserving 30 per cent of land and sea areas. As West and Brockington (2006, 609) argue, PAs have become ‘a way of thinking about the world, of viewing the world, and of acting on the world’. As such, this approach has been criticised for being too simplistic and too Western (Berkes, 2004). For a large number of people, the vision of conserving the wilderness or wild landscapes is nothing more than a Euro-American urban vision that has been imposed through colonial and neo-colonial regimes (Pascual et al., 2021). In this logic we can include the ‘new’ ecomodernism or ecopragmatism approach (Asafu-Adjave et al., 2015), which argues for a good Anthropocene through better technology, urbanisation, and the decoupling of people from nature. Once

Biodiversity conservation  27

again, this interpretation returns to the old discourse of mythologising the ‘wilderness’, intensifying agriculture for land-sparing with a green revolution and synthetic fertilisers’ logic, and market economic efficiency for decoupling human well-being. This Western conservation approach does not respond to the realities of local communities, in many cases Indigenous, who practice different society–nature relations with a multiplicity of values that are different from those of the West. A common feature of these other ways of conceiving society–nature relations is that people are not external to biodiversity, isolated from nature. It may seem obvious, but recognising the plurality of values is fundamental to making visible those values that are not taken into account in conservation decisions, but that are key when it comes to understanding biodiversity and its relationship with human well-being. Arguments such as the above have led biodiversity conservation to broaden its conceptualisations, approaches, methods, and tools. For several decades, local conservation initiatives have been promoted and consolidated; these are initiatives that go beyond the vertical imposition of a PA where the legal, cultural, social, and economic contexts of the local population are, most commonly, not considered. Local conservation initiatives tend to be less restrictive to local livelihoods and often incorporate management mechanisms that are more representative of local institutions and their society–nature relationships. However, regardless of the conservation model, it is necessary to return to the argument made by Alcorn (1994), that proof of biodiversity conservation success needs to be valued—independently of the context and the knowledge and value system. Just because local conservation initiatives are implemented by local communities and are based on their traditional natural resource management strategies does not mean that biodiversity is being conserved. In this regard, one of the central aspects of biodiversity conservation is that decision-making must be based on evidence. Of course, if we are talking about biodiversity, per se, biological evidence is what will allow us to know if the actions we are implementing are going in the right direction. However, as already argued, conservation is a socio-political process in which insti-

tutions, organisations, enterprises, the economic system, and so on play a crucial role in the possible success that can be achieved. Therefore, in all aspects (legal, social, political, economic, philosophical, and ethical), evidence is needed for policy design and decision-making. The debate is not whether that evidence has to be exclusively scientific or whether it can come from other knowledge and value systems: what we need is relevant information on which to base decisions, and these decisions need to be constantly evaluated and challenged. This does not mean that conservation is a puzzle-solving exercise and that the more evidence we generate the more efficient we will be at conserving biodiversity. Regardless of the context, in the conservation arena we are always dealing with high stakes and high levels of uncertainty—and values are always in dispute (Francis and Goodman, 2010). Therefore, by its very nature, conservation is an inexhaustible source of conflicts.

5.3

Conservation conflicts

In this final section, I show the different interpretations of biodiversity conservation conflicts, because managing conflicts is arguably one of the most important tools for fulfilling conservation objectives (Castro and Nielsen, 2003). Conflicts are an inherent element of biodiversity conservation. Regardless of the conservation model and the instruments chosen to achieve the objectives, disputes are extremely common between actors (governments, local populations, NGOs, the business sector, etc.) about access to and the use of natural resources. Nowadays, it is also common to see the same actors fighting from the same trench against ambitious capital investment projects (mining, dam and highway construction, tourism and/or urbanisation developments) that threaten the health of people and biodiversity. Even though the conservation field has had decades of experience in researching conflicts, it is interesting to observe that there is no consensus about ‘the object of analysis itself: the conflict’ (Paz, 2014, 3). For example, ‘conflicts’ are often mistaken for ‘problems’, although many problems— such as the proliferation of invasive marine algae—do not represent a conflict situation. Conflicts are a characteristic of human society and develop in many forms (Jeong, 2008). It Eduardo García-Frapolli

28  Elgar encyclopedia of ecological economics

is often argued that conflicts arise because the people or institutions involved have differences or incompatibilities in their interests, values, power, perceptions, and objectives about something in particular. Nevertheless, although conflicts are based on differences, not all differences automatically become conflicts (Glasl, 1999). The traditional approach is what the literature defines as ‘human–wildlife conflicts’. Connover (2001) defines these conflicts as those that occur when the action of a human being has an adverse effect on wildlife, and vice versa. There are many examples of this type of conflict: conflicts about the predation of livestock by large carnivores (Bagchi and Mishra, 2006; Gillingham and Lee, 1999; Kissui, 2008); the negative effect of fauna on agricultural harvests (Hill, 2000); or the increase in encounters between humans and wildlife because of the rise in urbanisation and the reduction of habitat (Gehrt et al., 2009; Soulsbury and White, 2015). This traditional approach for understanding conservation conflicts has its origin in the natural sciences, particularly conservation biology. What can be observed in the research about conflicts between ‘humans and wildlife’ is that the approach draws on very concrete case studies; in other words, the case studies are based on particular experiences, and from these the understanding of conflict is built. Understanding from where and how these studies emerge makes it possible to see that the large majority lack a coherent social theory that can explain the conflict’s structure. Moreover, many studies lack a concrete description of how the researchers have interpreted the conflict—or how to tackle it. What has happened in recent years is that much research has incorporated theory and concepts from management or business administration for the sake of contributing to ‘conflict resolution’. In PAs, these conflicts are common and add another element to the confrontation between people and wildlife. Many local residents, as Dickman (2010) and Skogen et al. (2008) have noted, feel antipathy towards the targeted wild species because, in addition to negative impacts on their livelihoods caused by damage—for example, to crops—they feel that the conservation of these species is imposed by a more powerful urban elite in a domain of inequality and power imbalance. Eduardo García-Frapolli

Although many studies are still based on this human–wildlife approach, new methods for understanding conservation conflicts have emerged. These new approaches argue that what we understand as conflicts related to wildlife or biodiversity are, in reality, conflicts between people about wildlife or other aspects of biodiversity (Pooley et al., 2017; White et al., 2009). In other words, conservation conflicts arise when the interests of two or more parties compete for a specific aspect of biodiversity and when at least one of the parties perceives that its interests have been sacrificed at the expense of the interests of the other party (White et al., 2009). From this perspective, conflicts are always between at least two actors (human beings and/or institutions). This approach radically changes the interpretation of conflicts. Rather than being perceived as people versus wildlife, conflicts are understood as being between people about some specific aspect of nature. This implies that we should not understand so-called ‘human–wildlife conflicts’ as being about humans versus wildlife as such, because there is no human counterpart. White et al. (2009) called these ‘biodiversity conflicts’. Redpath et al. (2013), building on this understanding, worked with the concept of conflicts of conservation. Conservation conflicts are those that occur when people or institutions clash over differences regarding conservation objectives and when one of the parties asserts—or at least perceives—that their interests have been sacrificed at the expense of the interests of the other party. One of the advantages of understanding conflicts from this perspective is that what defines the existence of a conflict is whether an action or the behaviour of an actor diminishes or harms the other actor. For Escobar (2006), these conflicts are, in fact, a historical struggle that has existed for centuries for forests, biodiversity, food, water, rivers, and seas. In this sense, understanding conservation conflicts is about recognising the power that actors have in deciding what, how, and where to conserve. The difference between this conceptualisation of conflicts and the ones previously described is that it explores the politicisation of nature through conflicts, instead of naturalising the conflicts through environmental analysis (Le Billon, 2015).

Biodiversity conservation  29

This approach offers a different perspective for understanding the conflicts between people about wildlife or other aspects of biodiversity. It is done recognising that all human decisions—and therefore also conservation— are inherently political (Adams, 2015), that the relations between actors occur in a realm of power (Robbins, 2011), and that this power, at least currently, occurs in the framework of a global neoliberal economy that aims to generate surpluses (Arsel & Büscher, 2012). Escobar (2006) argues that what is actually occurring at present in many biodiverse regions is the transformation of diverse local economies—in part oriented to self-production and subsistence—to economies propelled by the market. This implies changing complex ecosystems into modern interpretations of nature (often plantations), and it also implies changing local cultures, based on place, into cultures that have to resemble the dominant modern culture more, with its individualistic and productive ethos and its orientation towards the market. It is clear that there are diverse approaches to understanding conservation conflicts, and that the way in which we address them has important implications both for their management and for policy recommendations. Defining a conflict as human–wildlife, or establishing an area as protected, or outlawing hunting in a determined place, are decisions that impose a basic set of beliefs that guide action: they impose a specific way of understanding nature and the production of knowledge, which is inseparable from social relationships of power (Bridge, 2004). Thus, as Moon and Blackman (2014) highlight, as practitioners of conservation, it is really important that we understand and recognise the principles and assumptions encrusted in our disciplines. Eduardo García-Frapolli

References

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Alcorn, J.B., 1994. Noble savage or noble state? Northern myths and southern realities in biodiversity conservation. Etnoecológica 2, 7–19. Arsel, M., Büscher, B., 2012. NatureTM Inc.: Changes and continuities in neoliberal conservation and market-based environmental policy. Development and Change 43, 53–78. https://​doi​ .org/​10​.1111/​j​.1467​-7660​.2012​.01752​.x Asafu-Adjaye, J., Blomquist, L., Brand, S., Defries, R., Ellis, E., Foreman, C., Keith, D., Lewis, M., Lynas, M., Nordhaus, T., Pielke, R., Pritzker, R., Roy, J., Sagoff, M., Shellenberger, M., Stone, R., Teague, P., 2015. An Ecomodernist Manifesto. Breakthrough Institute, Oakland, CA. Bagchi, S., Mishra, C., 2006. Living with large carnivores: predation on livestock by the snow leopard (Uncia uncia). Journal of Zoology 268, 217–24. https://​doi​.org/​10​.1111/​j​.1469​-7998​ .2005​.00030​.x Berkes, F., 2004. Rethinking community-based conservation. Conservation Biology 18, 621–30. https://​doi​.org/​10​.1111/​j​.1523​-1739​ .2004​.00077​.x Bridge, G., 2004. Mapping the bonanza: geographies of mining investment in an era of neoliberal reform. The Professional Geographer 56, 406–21. https://​doi​.org/​10​.1111/​j​.0033​-0124​ .2004​.05603009​.x Cardinale, B.J., Duffy, J.E., Gonzalez, A., Hooper, D.U., Perrings, C., Venail, P., Narwani, A., MacE, G.M., Tilman, D., Wardle, D.A., Kinzig, A.P., Daily, G.C., Loreau, M., Grace, J.B., Larigauderie, A., Srivastava, D.S., Naeem, S., 2012. Biodiversity loss and its impact on humanity. Nature 486, 59–67. https://​doi​.org/​ 10​.1038/​nature11148 Castro, A.P., Nielsen, E., 2003. Natural resource conflict management case studies: an analysis of power, participation and protected areas. Food and Agriculture Organization of the United Nations, Rome. Connover, M.R., 2001. Resolving Human–Wildlife Conflicts: The Science of Wildlife Damage Management. CRC Press, Boca Raton, FL. Convention on Biological Diversity (CBD), 1992. Convention on Biological Diversity. CBD, New York. Dickman, A.J., 2010. Complexities of conflict: The importance of considering social factors for effectively resolving human–wildlife conflict. doi​ Animal Conservation 13, 458–66. https://​ .org/​10​.1111/​j​.1469​-1795​.2010​.00368​.x Escobar, A., 2006. Difference and conflict in the struggle over natural resources: a political ecology framework. Development 49, 6–13. https://​doi​.org/​10​.1057/​palgrave​.development​ .1100267 Francis, R.A., Goodman, M.K., 2010. Post-normal science and the art of nature conservation.

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30  Elgar encyclopedia of ecological economics Journal for Nature Conservation 18, 89–105. https://​doi​.org/​10​.1016/​j​.jnc​.2009​.04​.002 Funtowicz, S.O., Ravetz, J., 1991. A new scientific methodology for global environmental issues, in: Costanza, R. (Ed.), Ecological Economics: The Science and Management of Sustainability. Columbia University Press, New York, pp. 137–52. Gaston, K.J., 1996. What is biodiversity?, in: Gaston, K.J. (Ed.), A Biology of Numbers and Difference. Blackwell Science Ltd., Oxford, pp. 1–18. Gehrt, S.D., Anchor, C., White, L.A., 2009. Home range and landscape use of coyotes in a metropolitan landscape: Conflict or coexistence? Journal of Mammalogy 90, 1045–57. https://​doi​ .org/​10​.1644/​08​-MAMM​-A​-277​.1 Gillingham, S., Lee, P.C., 1999. The impact of wildlife-related benefits on the conservation attitudes of local people around the Selous Game Reserve, Tanzania. Environmental Conservation 26, 218–28. https://​doi​.org/​10​ .1017/​S0376892999000302 Glasl, F., 1999. Confronting Conflict: A First Aid Kit for Handling Conflict. Hawthorn Press, Stroud, UK. Hill, C.M., 2000. Conflict of interest between people and baboons: Crop raiding in Uganda. International Journal of Primatology 21, 299–315. https://​doi​.org/​10​.1023/​A:​ 1005481605637 Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES), 2019. Global assessment report of the Intergovernmental Science-Policy Platform on Biodiversity. Bonn, Germany, IPBES Secretariat. Jeong, H.-W., 2008. Understanding Conflict and Conflict Analysis. SAGE, London. https://​doi​ .org/​10​.1017/​CBO9781107415324​.004 Kareiva, P.M., Marvier, M., Silliman, B.R., 2018. Effective Conservation Science: Data Not Dogma. Oxford University Press, Oxford. Kissui, B.M., 2008. Livestock predation by lions, leopards, spotted hyenas, and their vulnerability to retaliatory killing in the Maasai steppe, Tanzania. Animal Conservation 11, 422–32. https://​doi​.org/​10​.1111/​j​.1469​-1795​.2008​ .00199​.x Le Billon, P., 2015. Environmental conflict, in: Perreault, T., Bridge, G., McCarthy, J. (Eds.), The Routledge Handbook of Political Ecology. Routledge, London, pp. 598–608. Moon, K., Blackman, D., 2014. A guide to understanding social science research for natural scientists. Conservation Biology 28, 1167–77. https://​doi​.org/​10​.1111/​cobi​.12326 Otero, I., Farrell, K.N., Pueyo, S., Kallis, G., Kehoe, L., Haberl, H., Plutzar, C., Hobson, P., García-Márquez, J., Rodríguez-Labajos, B., Martin, J.L., Erb, K.H., Schindler, S.,

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Nielsen, J., Skorin, T., Settele, J., Essl, F., Gómez-Baggethun, E., Brotons, L., Rabitsch, W., Schneider, F., Pe’er, G., 2020. Biodiversity policy beyond economic growth. Conservation Letters 13, 1–18. https://​doi​.org/​10​.1111/​conl​ .12713 Pascual, U., Adams, W.M., Díaz, S., Lele, S., Mace, G.M., Turnhout, E., 2021. Biodiversity and the challenge of pluralism. Nature Sustainability 4, 567–72. https://​doi​.org/​10​.1038/​s41893​-021​ -00694​-7 Paz, M.F., 2014. Conflictos socioambientales en México: ¿Qué está en disputa?, in: Paz, M.F., Risdell, N. (Eds.), Conflictos, Conflictividades y Movilizaciones Socioambientales En México: Problemas Comunes, Lecturas Diversas. CRIM-UNAM y Miguel Ángel Porrúa, Cuernavaca, pp. 11–61. Pooley, S., Barua, M., Beinart, W., Dickman, A., Holmes, G., Lorimer, J., Loveridge, A.J., Macdonald, D.W., Marvin, G., Redpath, S., Sillero-Zubiri, C., Zimmermann, A., Milner-Gulland, E.J., 2017. An interdisciplinary review of current and future approaches to improving human–predator relations. Conservation Biology 31, 513–23. https://​doi​ .org/​10​.1111/​cobi​.12859 Redpath, S.M., Young, J., Evely, A., Adams, W.M., Sutherland, W.J., Whitehouse, A., Amar, A., Lambert, R.A., Linnell, J.D.C., Watt, A., Gutiérrez, R.J., 2013. Understanding and managing conservation conflicts. Trends in Ecology and Evolution 28, 100–109. https://​doi​.org/​10​ .1016/​j​.tree​.2012​.08​.021 Robbins, P., 2011. Political Ecology: A Critical Introduction. Blackwell Publishing, Malden, MO. Schulze, K., Knights, K., Coad, L., Geldmann, J., Leverington, F., Eassom, A., Marr, M., Butchart, S.H.M., Hockings, M., Burgess, N.D., 2018. An assessment of threats to terrestrial protected areas. Conservation Letters 11, 1–10. https://​doi​.org/​10​.1111/​conl​.12435 Shiva, V., 2000. Tomorrow’s Biodiversity. Thames & Hudson, London. Skogen, K., Mauz, I., Krange, O., 2008. Cry wolf! Narratives of wolf recovery in France and Norway. Rural Sociology 73, 105–133. https://​ doi​.org/​10​.1526/​003601108783575916 Soulé, M.E., 1985. What is conservation biology? Bioscience 35, 727–734. https://​doi​.org/​10​ .11647/​obp​.0177​.01 Soulsbury, C.D., White, P.C.L., 2015. Human– wildlife interactions in urban areas: A review of conflicts, benefits and opportunities. Wildlife Research 42, 541–53. https://​doi​.org/​10​.1071/​ WR14229 West, P., Brockington, D., 2006. An anthropological perspective on some unexpected consequences of protected areas. Conservation

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grated conceptual framework to understand biodiversity conflicts. Land Use Policy 26, 242–53. https://​doi​.org/​10​.1016/​j​.landusepol​ .2008​.03​.005

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6.

Bounded openness over natural information

discussions on DSI at COP15. Although controversy rages over the placeholder, efforts to vet alternative terms have been only half-hearted. The note puts “natural information” on the table. COP13 commissioned the first study on DSI, which was followed by four more at COP14. In the intersessional period, the Secretariat hosted an online discussion as well as several webinars on DSI (CBD, 2021b). As of this writing, a Google Scholar search of DSI generates 506 hits, and a Google search reveals some 41 600. In the note on DSI, the Executive Secretary invited the submission of views with a deadline of four weeks. Only 11 of 196 parties to the CBD and 17 of umpteen stakeholders responded. The invitation was, in essence, a two-part query: Does the third objective of Article 1 of the CBD, viz. “access to genetic resources” and “the fair and equitable sharing of benefits arising [from their] utilization” (Access and Benefit-Sharing [ABS] Capacity Development Initiative, 2019), apply to genetic resources once dematerialized? And if so, how? Users of genetic resources from the North have long argued that “material” in “genetic material” refers only to the tangible component in research and development (R&D). Japan repeats this position with vehemence, whereas Brazil argues that “material” has always included information (CBD, 2021a). More than 1600 databases now exist on dematerialized genetic resources (Rigden & Fernandez, 2019). Is biopiracy happening on a vast and almost unimaginable scale? Or is this all just for show? The jumping together of knowledge is fruitful, and the result merciless (Box 6.1). Natural information is the object of access for R&D, whether the genetic resource is dematerialized or not. From this reality flows “bounded openness over natural information” (Vogel, 2018b).

Introduction

The naturalist E.O. Wilson (1998) was fond of quoting the Chinese saying that “the first step to wisdom is getting things by their right names” (p. 4). This simple advice would help realize the objectives of the 1992 Convention on Biological Diversity (CBD), which are conservation, sustainable use, and the fair and equitable sharing of benefits arising from their utilization (CBD, 1992). The three intertwined objectives fall squarely within the domain of ecological economics. What is the right name for the biological object of utilization? The answer to be explored here is “natural information.” And what is the right name for the modality that will govern the fair and equitable sharing of benefits from utilization? The answer is “bounded openness.” Modification of “natural information” with “bounded openness” generates a powerful abstraction not only for the CBD but also more broadly for science policy and epistemology. The neologism is, however, only suggestive of the intended meaning; “bounded openness over natural information” requires unpacking. Clarity emerges through uniting disparate fields and entertaining questions of provenance. Explanation may seem roundabout as interpretation requires diligence. This “‘jumping together’ of knowledge,” or consilience (Wilson, 1998, p. 8), coheres with the 1969 Vienna Convention on the Law of Treaties, specifically Article 31 “General Rule of Interpretation” and Article 32 “Supplementary Means of Interpretation.”

In media res

“Bounded openness over natural information” appears en passant in the note by the Executive Secretary of the CBD of September 3, 2021, titled “Digital Sequence Information on Genetic Resources” (CBD, 2021a). The expression in the title is known by the acronym DSI and became the placeholder for dematerialized genetic resources at the 13th Conference of the Parties to the CBD (COP13). The 22-page note hopes to orient

BOX 6.1 A PESKY CAT AND PET THEORIES “A synthesis of economics and chemistry invites a thought experiment [á la Schrödinger’s Cat]: denature the material transferred in an MTA [Material Transfer Agreement] and then perform R&D. By the First Law of Thermodynamics, the sample will have retained all of its matter, but by the

32

Bounded openness over natural information  33 Second, much of the associated information will be lost. One deduces that the “material” in an MTA should not be interpreted as matter, though legally it is. The value lies in the information as the matter would still be there upon denaturation. A corollary exists: a sample returned in a pristine state to the property owner can also have lost all value in exchange, similar to denaturation, as the owner no longer has any leverage over granting access to the information therein.” (Sociedad Peruana de Derecho Ambiental [SPDA], 2021)   “The law that entropy always increases—the Second Law of Thermodynamics—holds, I think, the supreme position among the laws of Nature. If someone points out to you that your pet theory of the universe is in disagreement with Maxwell’s equations—then so much the worse for Maxwell’s equations. If it is found to be contradicted by observation— well, these experimentalists do bungle things sometimes. But if your theory is found to be against the Second Law of Thermodynamics, I can give you no hope; there is nothing for it but to collapse in deepest humiliation.” (Sir Arthur Eddington, The Nature of the Physical World, 1915)

Provenance

“Bounded openness” first appears with “natural information” in “The Economics of Information Studiously Ignored in the Nagoya Protocol” (Vogel et al., 2011). The title reflects a research stream that was already two-decades old at the time of its publication (Stone, 1995; Swanson, 1992; Swanson et al., 1994; Vogel, 1991, 1992, 1994). The terms “bounded openness” and “natural information” were still not, however, conjoined. The fusion occurred two years later in an online discussion hosted by the Secretariat to the CBD, with some 140 participants (CBD, 2013; Vogel et al., 2018). From 2013 forward, “bounded openness over natural information” would become synonymous with a proposed modality for the Global Multilateral Benefit-Sharing Mechanism (GMBSM), which is the title of Article 10 of the 2010 Nagoya Protocol (NP; Ruiz Muller, 2015).

BOX 6.2 STEPWISE INSTITUTIONALIZATION OF “BOUNDED OPENNESS OVER NATURAL INFORMATION” 1. Unencumbered access: Parties reconfigure ABS Competent National Authorities concomitant with the ratification of an amendment to the NP that enables the GMBSM; 2. Disclosure: Intellectual property institutions adapt application procedures to include a Yes/No query or other method, regarding the utilization of natural information; 3. Royalties upon commercialization: On net sales disclosed annually, the GMBSM charges a percentage, the negotiation of which is foreseen in the aforementioned amendment to the NP, with income held in escrow; 4. First-round cost estimates: From the royalties in escrow, GMBSM engages (a) molecular biologists to estimate the costs of determining homologous diffusion of natural information as well as diffusion in distinct lineages, should convergent evolution be evidenced, and (b) ecologists to model the habitat of species so identified; 5. Financial threshold of estimates never met: Royalties in escrow revert to GMBSM upon expiry of intellectual property associated with the utilization; 6. Financial threshold met: From the royalties in escrow, GMBSM engages biologists and ecologists to conduct aforementioned studies; 7. Second-round estimates: From the royalties in escrow, GMBSM engages field biologists to estimate the costs of “ground truthing” (i.e., verification of the presence of species in the habitat[s]); 8. Financial threshold never met: Royalties in escrow revert to the GMBSM upon expiry of the intellectual property associated with the utilization; 9. Financial threshold met: GMBSM authorizes studies and distributes royalties annually according to share of habitat positively ground-truthed and opens a temporal window to evaluate new claims for, or challenges to, Provider status [for marine genetic resources, distribution according to relative performance in surpassing agreed reductions of CO2e emissions]; 10. Expiry of associated intellectual property: The claims of providers also expire and natural information enters public domain. (Vogel, 2018a)

Joseph Henry Vogel, María Eugenia Santori-Aymat, Óscar Tomaiconza et al.

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Despite a lively exchange in the 2013 online discussion, “bounded openness over natural information” remained taboo in COP12 to COP14, just as “natural information” had been taboo in the previous COPs. Recognition of that status can be inferred from the closing sentence of the report by the 2018 Ad Hoc Technical Expert Group (AHTEG) on DSI “‘bounded openness over natural information’ may merit consideration” (CBD, 2018). The suggestion to break the taboo became a devoir with the First Global Dialogue on DSI, held November 6–8, 2019, in Pretoria, South Africa. The 65 participants from 27 countries were advised that “What should not happen in this dialogue . . . Taboos and restriction in discussions.” “Bounded openness” was cited twice in the Report of the First Global Dialogue (ABS Capacity Development Initiative, 2019). The 2021 Note from the Executive Secretary is the highest-profile medium to break the natural-information taboo. However, any taboo three-decades in the making does not disappear with an utterance or two. Overcoming a cultivated resilience requires an understanding of causation. For “bounded openness over natural information,” disparate disciplines must jump together: ● Historians date the identification of taboos in Western thought with Captain Cook’s voyage to Tonga, Polynesia, in 1777; ● Anthropologists observe the cohesive effect of taboos on tribes over recorded history; ● Sociobiologists transcend anthropocentrism and classify taboo as an expression of “nested [dominance] hierarchies” (Wilson, 1975, p. 287), which are common across Animalia; ● Psychologists explain the resilience of the natural-information taboo as positive reinforcement from official discussions in COP1 to COP14; ● Economists chime in with the “fallacy of sunk costs.” Resources misallocated biased decisions in favor of continued misallocation. Other economists perceive a principal–agent problem. The principals are the uninformed citizenry, and the agents are heads of delegations and bureaucrats. Admission of error hazards accountability;

● Entomologists wink “the more elaborate and expensive the nest is in energy and time, the greater the fierceness of the ants that defend it” (Wilson, 2012, p. 130, emphasis in original); ● Lawyers inside the delegations wield the trump card, or at least think they do. Rather than invoke stare decisis and appear lawyerly, they deploy folksy language; ● Cognitive linguists recognize the homilies as positive mental framing (Lakoff, 1987): the ship has sailed; the horse is out of the barn; bilateral ABS is baked into the system. Delegates sigh and tastebuds fire; ● Lawyers outside the delegations interject that stare decisis does not hold if a decision is fundamentally wrong; ● Wonks perceive a natural experiment in the Note. Assuming the causation runs dominance → fallacy → taboo → extinction, the submissions can be formalized as an easily testable hypothesis: ● H0: Commission of the sunk-cost fallacy, the taboo over natural information and nested dominance hierarchies will not frustrate recognition of the economic solution for ABS; ● Ha: Commission of the sunk-cost fallacy, the taboo over natural information and nested dominance hierarchies will frustrate recognition of the economic solution for ABS; ● Data: 27 of 28 submissions do not recognize “bounded openness over natural information” cited in the Note. The null hypothesis is rejected. ● Sociobiologists delight. Eusocial behavior to respect taboos, commit fallacies, and submit to dominance hierarchies has all the earmarks of instinct; recall Lotka’s ● Thermodynamicists Principle (jokingly called the Fourth Law). Evolved dominance hierarchies throughout Animalia increase entropy (Vogel, 1988). The causal chain is amended leftward: thermodynamics → dominance → fallacy → taboo → extinction; ● Social activists, inspired by a 19th-century German philosopher who often goes unnamed, perceive that the problem is The System. They sense urgency; ● Artists create polemical and inspirational works, but eschew the details of just how to change the system;

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● Ecocritics argue that limits, binding over generations, are the change most needed (Gomides, 2006). None is more easily comprehensible than half-Earth (Wilson, 2016); bounded openness over natural information eases the transition (Vogel, 2018b). Such “mutual coercion, mutually agreed upon” (Hardin, 1968) assumes engagement (Vogel, 2008, 2010), where art becomes indispensable to pierce the public sphere (Vogel, 2018c); ● Economists revel in the abstraction. Inasmuch as continued evolution of all species is the objective of conservation, society must contemplate that autocracy will return in the future of life. Without the rule of law, how will biodiversity survive in 100, 1000, or 10 000 years? Evolutionary time is measured in thousands and tens of thousands of years. The abstraction seems ethereal: perhaps only cultivation of a taboo over relaxing half-Earth can transcend political retrogression; and finally ● Historians and anthropologists perceive an irony. Cosmovision is a web of limits and typical of pre-literate societies. The causal chain is amended rightward: thermodynamics → dominance → fallacy → taboo → extinction → taboo. ● Consilience obtains. Each discipline is viewed as a layer of analysis whose judicious selection and synthesis affords Gestalt. Reduction is guided by anticipated construction. Five words are contained in “bounded openness over natural information.” Analysis partitions the neologism and decomposes each pair of words. Consideration of the whole is the last step.

What is “bounded openness”?

“Bounded openness” first appears in the academic literature with distinct meanings for different fields, including psychology (Brandt et al., 2015) and religion (Jones, 2001). We do not consider those meanings as such consideration risks equivocation. The desired layer of analysis is political economy. Christopher May (2010) elaborated the relevance of “bounded openness” in The Global Political Economy of Intellectual Property

but did not define it. The COP experience with “material” demonstrates the folly of leaving a key term undefined. In a forum populated by lawyers, “bounded openness” must not only be defined but also contain exclusionary and inclusionary criteria. The SPDA (2016a) offered the following definition, inspired by the work of May: “Bounded openness: Legal enclosures which default to, yet depart, from res nullius to the extent the departures enhance efficiency and equity, which must be balanced when in conflict.” Lawyers may pause in disbelief. Nudge, nudge: Isn’t “bounded openness” what intellectual property already does? Is this just old wine in a new bottle (Vogel et al., 2021)? Such skepticism is welcome. The mindset of bounded openness is weighted more toward the openness than the bounds. As a modality for the management of information, bounded openness is not confined to intellectual property. The inclusiveness invokes equal treatment for that which is natural.

What is “openness”?

The online Merriam-Webster Dictionary lists 20 meanings. Only the seventh and the eleventh seem appropriate for ABS: ● presenting no obstacle to passage or view: not enclosed, obstructed, or filled with objects; . . . ● characterized by ready accessibility and usually generous attitude. The context of “bounded openness over natural information” is “open access,” which has been defined as “free, unrestricted online access to research outputs such as journals, articles and books . . . open to all, with no access fees” (Springer Nature, 2021). Any terms and conditions, no matter how minimal, are bounds. Such interpretation runs counter to “Open Science,” which is the UNESCO classification where some restrictions are permissible, contrary to how “open” is generally understood. Replacement of “Open Science” with, say, “Science Open yet Bounded” would probably face the same hurdles encountered with “natural information” in the COP, viz. nested dominance hierarchies, the sunk-cost fallacy, and taboo.

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What is “bounded”?

Merriam-Webster offers three interrelated meanings of the past participle of “to bound.” The second seems the most appropriate: “to set limits to (confine).” Interpretation of “bounded” is reminiscent of the First Law of Thermodynamics. The general public hears “bounded” and may think of inviolable physical law. The resulting mental frame is amenable to half-Earth. When the economist hears “bounded,” they may anticipate that the next word will be “rationality.” The polymath Herbert A. Simon won the 1978 Nobel Memorial Prize in Economics for exploring “bounded rationality” in public administration. If Keynes is the Darwin of Economics, then Simon is the Mendel. Like the Czech monk, the full impact of Simon’s genius has been lagged two human generations. Recognition of cognitive limitations and capacities undergird modern “behavioral economics.” Economists are mathematically inclined and know that one limit can justify another (Lipsey & Lancaster, 1956). “Bounded openness” puts the economist in a mental frame amenable to half-Earth.

What is “natural information”?

The term first appeared in reference to genetic resources in a newsletter from the Centre for International Research on Communications and Information Technologies (Vogel, 1991). Its meaning would be considered self-evident in a research stream that spanned decades (Ruiz Muller, 2015). Such luxury ended in 2016 when “natural information” figured into the title of a Side Event at COP13 (SPDA, 2016b). Two complementary definitions modified by biotic or abiotic appeared in the third version of a proposed amendment to the NP and were accompanied by their complement: ● Natural Information (biotic): Any unintentional distinction, non-uniformity or difference extracted from matter that is living or was once alive. ● Natural Information (abiotic): Complement of Natural Information (biotic) with respect to that which is not living and was never alive. ● Artificial Information: Any human-made distinction, non-uniformity or difference that is intentional (SPDA, 2021).

What is “information”?

“Information is decrease in uncertainty” (Schneider, 2018). This terse sentence derives from the Boltzmann equation of thermodynamics or the isomorphic Shannon equation of information theory. The mathematics of the formulas is daunting. Suffice it to say here that the Second Law of Thermodynamics not only implies limits but also the identity of the object of access for R&D in biotechnology.

What is “natural”?

The cynic will answer “nothing—we live in the Anthropocene!” The non-cynic answers that natural information is distinguished from artificial through intent. For example, variants to the COVID-19 virus may result from not wearing masks, not social distancing, and not being vaccinated. The unintended variant Delta is natural information. An intended variant, G-d forbid, would be artificial information.

What is “over”?

Prepositions are the bane of any language. Aesthetics enters when choice is unclear. Why is “over” preferable to “across,” “above,” “through,” “in,” or “for”? “Over” rather than “across” implies that “bounded openness” lies at a higher plane. “Over” rather than “above” implies that “bounded openness” spans a vast domain of “natural information.” Other prepositions, such as “through,” “in,” and “for,” invoke inappropriate nuances and do not work aesthetically.

What is “bounded openness over natural information”?

Three meanings can be adduced from the above discussion: 1. Legal enclosures over any unintentional distinction, non-uniformity, or difference extracted from matter that is living or was once alive that default to, yet depart from, res nullius to the extent the departures enhance efficiency and equity, which must be balanced when in conflict; 2. The proposed modality of the GMBSM of the NP; 3. The principal means for half-Earth

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Bounded openness over natural information  37

Conclusion

History humbles the economist. New abstractions are seldom wholly original. The classical economist David Ricardo (1772–1823) fought tirelessly in the House of Commons to repeal the tariffs on grains. The protectionism of the Corn Laws resulted in high prices and “rents”—Ricardo’s abstraction for income unjustly distributed through the market mechanism. Factory workers had to eat; wages rose to cover the food budget. In essence, landlords were reaping the greatest benefit from industrialization. England could not fully industrialize until the repeal of the Corn Laws in 1846—roughly a generation after the death of Ricardo. Similar deductive reasoning applies to biological diversity in this the Fourth Industrial Revolution. Distinct premises will render different conclusions. Unlike grains, genetic resources are intangible and best conceived as natural information. The costs of reproduction are insignificant: a few kilos of dried leaves, a Ziploc bag of scooped soil, or a downloaded genetic sequence. “Fair and equitable” ABS should be interpreted as a limit that enables rents. Karl Marx wrote in Grundrisse that industry converts any limit into a barrier, which it then strives to circumvent or transcend 408). Users of genetic (1857–61/1993, p.  resources are living up to that insight, shamelessly. Under bilateralism, royalty percentages are as low as 0.1 percent in the most biodiverse country on the planet (Brazil, 2015). Because rents are effectively eliminated, Providers do not facilitate access. Thinking of Adam Smith, why should they? Imbecility triumphs whenever mega-diverse Providers intone sovereignty against multilateralism. Bounded openness over natural information is the correction, long overdue. Only once corrected will half-Earth ever emerge.

Acknowledgments

We would like to thank the Sociedad Peruana de Derecho Ambiental for sponsorship of open-access publication and for posting translations of this and related works. We have benefited greatly from the discerning comments of Newton C. Fawcett, Stanley P. Kowalski, Omar Oduardo-Sierra, Klaus Angerer, Camilo Gomides, and Manuel Ruiz Muller.

This entry is available for free as Open Access from the individual product page at www​ .elgaronline​.com under a Creative Commons Attribution-NonCommercial-NoDerivatives 4.0 Unported (https://​creativecommons​.org/​ licenses/​by​-nc​-nd/​4​.0/​) license. Joseph Henry Vogel, María Eugenia Santori-Aymat, Óscar Tomaiconza, Bryan Steven Cortés-Lumbi, and Miguel Fernández-Maldonado

References

Access and Benefit-Sharing Capacity Development Initiative. (2019). The report of the First Global Dialogue on DSI. https://​ www​.abs​-biotrade​.info/​fileadmin/​Downloads/​ EVENT​%20REPORTS/​2019/​20191​-ABS​-I​-1st​ -DSI​-Dialogue​-South​-Africa​.pdf Brandt, M. J., Chambers, J. R., Crawford, J. T., . . . Reyna, C. (2015). Bounded openness: the effect of openness to experience on intolerance is moderated by target group conventionality. Journal of Personality and Social Psychology 109(3), 549–68. Brazil: Law No. 13.123. (2015). Article 20. http://​www​.wipo​.int/​edocs/​lexdocs/​laws/​pt/​br/​ br161pt​.pdf Convention on Biological Diversity (CBD). (1992). Text of the convention. https://​www​.cbd​ .int/​convention/​text/​ CBD. (2013, April 28). ABSCH: Discussion Forum Article 10 Nagoya Protocol. Comment #5148. https://​absch​.cbd​.int/​es/​portals/​forums/​ article​-10​-forum/​thread/​5079 CBD. (2018). Report of the Ad Hoc Technical Expert Group on Digital Sequence Information on Genetic Resources, CBD/DSI/ AHTEG/2018/1/4/. https://​www​.cbd​.int/​doc/​ c/​4f53/​a660/​20273​cadac31378​7b058a7b6/​dsi​ -ahteg​-2018​-01​-04​-en​.pdf CBD. (2021a). Note from Executive Secretary. Submission of views and new information on policy approaches, options or modalities for digital sequence information on genetic resources. SCBD/NPU/TS/CGA/AC/89861. https://​www​.cbd​.int/​conferences/​post2020/​ submissions/​2021​-063 CBD. (2021b). Webinar series on Digital Sequence Information on genetic resources. https://​www​ .cbd​.int/​article/​dsi​-webinar​-series​-2020 Hardin, G. (1968). The tragedy of the commons. Science 162(3859), 1243–8. Gomides, C. (2006). Putting a new definition of ecocriticism to the test: the case of The Burning Season, a film (mal)adaptation. ISLE 13(1), 13–23. Jones, S. (2001). Bounded openness: postmodernism, feminism, and the church today.

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38  Elgar encyclopedia of ecological economics Interpretation: A Journal of Church and Theology 55(1), 49–59. Lakoff, G. (1987). Women, Fire and Dangerous Things. Chicago: University of Chicago Press. Lipsey, R. G., & Lancaster, K. (1956). The general theory of second best. Review of Economic Studies 24(1), 11–32. Marx, K. (1993 [1857–61]). Grundrisse (M. Nicolaus, Trans.). Penguin Books. https://​ www​.marxists​.org/​archive/​marx/​works/​1857/​ grundrisse/​ May, C. (2010). The Global Political Economy of Intellectual Property Rights, 2nd ed. London: Routledge. Rigden, D. J., & Fernandez, X. M. (2019). The 26th annual nucleic acids research database issue and molecular biology database collection. Nucleic Acids R 47(D1), D1–D7. Ruiz, M. (2015). Genetic Resources as Natural Information. London: Routledge. Schneider, T. (2018). Information theory primer. http://​users​.fred​.net/​tds/​lab/​papers/​primer/​ primer​.pdf SPDA. (2016b). New approaches to access and benefit sharing: the case for bounded openness and natural information. Side Event at COP13 to CBD. http://​www​.actualidadambiental​.pe/​wp​ -content/​uploads/​2017/​12/​Transcrip​tSideEvent​ COP13Bound​edOpenness​.pdf SPDA. (2021). Fairness, equity and efficiency for the Convention on Biological Diversity and the Nagoya Protocol: analysis of a rodent, a snail, a sponge and a virus. Eschborn, Germany: The ABS Capacity Development Initiative. https://​ spda​.org​.pe/​?wpfb​_dl​=​4662 Springer Nature. (2021). What is open access? https://​www​.springernature​.com/​gp/​open​ -research/​about/​what​-is​-open​-access Stone, C. D. (1995). What to do about biodiversity, property rights, public goods and the Earth’s biological riches. Southern California Law Review 68(3), 577–605. Swanson, T. M. (1992). The Economics of the Biodiversity Convention. Norwich: CSERGE, School of Environmental Sciences, University of East Anglia. Swanson, T. M., Pearce. D. W., & Cervigni, R. (1994). The Appropriation of the Benefits of Plant Genetic Resources for Agriculture: An Economic Analysis of the Alternative Mechanism for Biodiversity Conservation. Rome: Secretariat of the FAO Commission on Plant Genetic Resources. Vogel, J. H. (1988). Evolution as an entropy-driven process: an economic model. Systems Research 5(4), 299–312. Vogel, J. H. (1991). The intellectual property of natural and artificial information. CIRCIT Newsletter. Melbourne: Centre for International

Research on Communication and Information Technologies. Vogel, J. H. (1992). Privatisation as a Conservation Policy. Melbourne: CIRCIT. Vogel, J. H. (1994). Genes for Sale. New York: Oxford University Press. Vogel, J. H. (2008). Ecocriticism as an economic school of thought: Woody Allen’s Match Point as exemplary. OMETECA Science and Humanities 12, 105–19. Vogel, J. H. (ed). (2010). The Museum of Bioprospecting, Intellectual Property and the Public Domain. London: Anthem Press. Vogel, J. H. (2018a, 11 May). The global multilateral benefit-sharing mechanism: where will be the Bretton Woods of the 21st century? Intellectual Property Watch/International Policy News. Vogel, J. H. (2018b, September 7). Not just a matter of matter: “the way forward” for the UNCBD, NP and half-Earth. Inside Views. Intellectual Property Watch/International IP Policy News. http://​www​.ip​-watch​.org/​2018/​ 09/​07/​not​-just​-matter​-matter​-way​-forward​ -uncbd​-np​-half​-earth/​ Vogel, J. H. (2018c). Wind blowin’ in the [brackets]: tribute to the Convention on Biological www​ Diversity [Video]. YouTube. https://​ .youtube​.com/​watch​?v​=​rYJx​-J31Op8 Vogel, J. H., Álvarez-Berríos, N. Quiñones-Vilche, N., Medina-Muñiz, J. L., Pérez-Montes, D., Arocho-Montes, A. I., . . . Santiago-Rios, J. (2011). The economics of information: studiously ignored in the Nagoya Protocol on access and benefit sharing. Law Environment and Development Journal 7(1), 52–65. http://​ www​.lead​-journal​.org/​content/​11052​.pdf Vogel, J. H., Angerer, K., Ruiz Muller, M., & Oduardo-Sierra, O. (2018). Bounded openness as the global multilateral benefit-sharing mechanism for the Nagoya Protocol. In C. R. McManis & B. Ong (eds.), Routledge Handbook on Biodiversity and the Law. London: Routledge, pp. 377–94. Vogel, J. H., Ruiz Muller, M., Angerer, K., Delgado Gutiérrez, D., & Gálvez Ballón. A. (2021). Bounded openness: a robust modality of access to genetic resources and the sharing of benefits. Plants, People, Planet 4(1), 13–22. https://​doi​.org/​10​.1002/​ppp3​.10239 Wilson, E. O. (1975). Sociobiology. Cambridge, MA: Harvard University Press. Wilson, E. O. (1998). Consilience. New York: Alfred P. Knopf. Wilson, E. O. (2012). The Social Conquest of Earth. New York: W. W. Norton. Wilson, E. O. (2016). Half-Earth. New York: W. W. Norton.

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7. Bounded rationality

myopic decisions. Bounded rationality can be a reason why traditional policy instruments, such as different types of economic incentives, can prove to be less effective than what is expected when assuming a rational calculation of costs and benefits (Handgraaf et al., 2013). Bounded rationality also gives rise to alternative instruments that “exploit” these human tendencies. These include some types of “nudges” (Schubert, 2017), in particular, so-called green defaults. Due to limited cognitive capacities and search costs, humans tend to stick with a default option. This is why, for example, making green electricity tariffs the default option shifts choices considerably to such tariffs (Kaiser et al., 2020). Furthermore, some ecological economists tend to be critical (and call for regulation) of commercial advertising, which exploits consumers’ bounded rationality (or social sensitivity), for example, through “green” advertising (Gsottbauer and van den Bergh, 2014). Research shows that adding green-labeled products to a shopping basket creates the illusion among consumers that the overall carbon footprint of the basket is lower than before (Holmgren et al., 2018). Finally, sustainable consumption and pro-environmental behavior are also hindered by another institutional factor, namely, lack of discretionary time. This lack, in turn, may increase people’s bounded rationality, preventing them to act upon their environmental preferences (Chai et al., 2015). The “rebound effect” is another major topic in ecological economics (Freire, Chapter 78, this volume). It refers to energy efficiency gains resulting from the adoption of a new technology or practice that are less than expected, given that the more efficient technology is used more, or the savings are spent on other energy-intensive products. Bounded rationality tends to increase direct and indirect rebound effects (Exadaktylos and van den Bergh, 2021). For example, some consumers simply are not aware that they are using a newly adopted technology more intensively. Others mentally classify energy savings as pure gains, keeping a mental account of these separate from the adoption outlays. This can then encourage re-spending of such perceived savings on other energy-intensive goods and services. “Systems” are important concepts in ecological economics as genuine sustainable solutions require that indirect or systemic

7.1 Introduction

Bounded rationality refers to the tendency of consumers, firms, and other economic actors to use mental shortcuts, resulting in non-optimal decisions. Bounded rationality is a response to the assumption of “perfect” rationality used in neoclassical economics. The latter refers to the idea that individuals make individually optimal decisions through a rational computation and comparison of expected utility derived from all available decision options. Instead of perfect calculations to achieve optimal outcomes, people use shortcuts and heuristics to arrive at satisfactory outcomes. This is due to a lack of cognitive abilities or simply a lack of time. The use of heuristics can lead to systematic biases. For example, when faced with a risky choice between a potential equal-sized loss and gain, people tend to respond more strongly to the loss (Kahneman, 2003). Bounded rationality is one of the central themes in the larger field of behavioral economics (Mullainathan and Thaler, 2000). While the term “bounded rationality” first appeared in an article in the journal Ecological Economics in 1990 (Tisdell, 1990), most published studies employing the term appear from the year 2000 on. Important articles published in this journal highlighting the relevance of bounded rationality ­– largely in the context of a broader set of behavioral issues – include van den Bergh et al. (2000) and Venkatachalam (2008). In the following, several applications of bounded rationality to research in ecological economics are presented.

7.2 Applications to ecological economics

The field in which bounded rationality probably plays the most important role is the assessment of policy instruments. Different types of bounded rationality have been integrated in economic models, notably in agent-based models (Castro et al., 2020). This involves boundedly rational behavior of consumers in studies of technology diffusion or energy conservation. It also covers models of firms’ boundedly rational behavior, such as having imperfect information or making 39

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effects, such as leakage, rebound, or problem shifting, are limited. However, research shows that many people lack the capacity for systemic thinking about complex environmental problems (Lezak and Thibodeau, 2016). Specifically, people scoring low on systemic thinking skills are less likely to ascribe monetary and non-monetary value to the environment (Lezak and Thibodeau, 2016). Of course, it is possible to learn systems thinking, just like it is possible to debias certain biases. One of the central ideas in ecological economics is to question whether continuous economic (gross domestic product [GDP]) growth on a finite planet is possible and desirable (Victor, 2010). There are not only environmental but also social limits to growth: research indicates no link between GDP and life satisfaction beyond a certain threshold (Easterlin, 1974), hence the discussions about alternative (well-being) indicators, including that “direct measures of experienced utility become particularly relevant in a context of bounded rational” (Kahneman et al., 2004). A distinct subset of the literature on growth/ environment has examined to what extent and why people support the economic growth paradigm (Tomaselli et al., 2019). One potential reason of support for the policy objective of growth is a lack of understanding of the characteristics of growth (van den Bergh, 2009). For example, research shows that many people do not know how economic growth – that is, GDP – is conventionally measured (Walstad, 1997). Most people in rich countries also overestimate present GDP growth rates (Drews et al., 2018). Perhaps most importantly, people have a hard time making sense of the dynamic effects of economic growth: when individuals are given a certain GDP growth rate and are asked to estimate the overall rise in national income over a certain time period, the vast majority significantly underestimates the increase in GDP (Christandl and Fetchenhauer, 2009). This has been called the “exponential growth bias” and may be one expression of bounded rationality contributing to the underestimation of the environmental impacts of continuous economic growth. In addition, when discussing alternatives to the growth paradigm, such as degrowth, research has highlighted various cognitive limitations of individuals that need to be taken into account to communicate such Stefan Drews

proposals more effectively (Drews and Antal, 2016). Due to missing markets, the contingent valuation method presents survey respondents an environmental good and asks them how much they are willing to pay for it, or how much they are willing to accept in compensation for its loss. Notwithstanding doubts on philosophical grounds (Sagoff, 1998), this method has been applied in many studies published in Ecological Economics. Methodological concerns about the method also involve bounded rationality, specifically, to what extent the latter can bias responses derived from contingent valuation. A study by Frör (2008) measures a form of bounded rationality, namely, individuals’ confidence to use intuitive-experiential or rational-analytical reasoning. The first type of reasoning was negatively, and the second type positively, associated with the willingness to pay for the improvement of the quality of household tap water. To conclude, bounded rationality has been successfully integrated in various research topics of ecological economics. Its integration is part of a larger attempt to provide better founded insights about causes and solutions to environmental problems through a more realistic representation of human behavior.

Acknowledgments

I am grateful to Jeroen van den Bergh for useful comments on a previous version of this chapter. Stefan Drews

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Castro, J., Drews, S., Exadaktylos, F., Foramitti, J., Klein, F., Konc, T., Savin, I., van den Bergh, J., 2020. A review of agent-based modeling of climate-energy policy. WIREs Climate Change 11, e647. https://​doi​.org/​10​.1002/​wcc​.647 Chai, A., Bradley, G., Lo, A., Reser, J., 2015. What time to adapt? The role of discretionary time in sustaining the climate change value– action gap. Ecological Economics 116, 95–107. https://​doi​.org/​10​.1016/​j​.ecolecon​.2015​.04​.013 Christandl, F., Fetchenhauer, D., 2009. How laypeople and experts misperceive the effect of economic growth. Journal of Economic Psychology 30, 381–92. https://​doi​.org/​10​ .1016/​j​.joep​.2009​.01​.002 Drews, S., Antal, M., 2016. Degrowth: A “missile word” that backfires? Ecological Economics

Bounded rationality  41 126, 182–7. https://​doi​.org/​10​.1016/​j​.ecolecon​ .2016​.04​.001 Drews, S., Antal, M., van den Bergh, J.C.J.M., 2018. Challenges in assessing public opinion on economic growth versus environment: Considering European and US data. Ecological Economics 146, 265–72. https://​doi​.org/​10​ .1016/​j​.ecolecon​.2017​.11​.006 Easterlin, R.A., 1974. Does economic growth improve the human lot? Some empirical evidence. Nations and Households in Econonomic Growth 89, 89–125. Exadaktylos, F., van den Bergh, J., 2021. Energy-related behaviour and rebound when rationality, self-interest and willpower are limited. Nature Energy 6, 1104–13. https://​doi​ .org/​10​.1038/​s41560​-021​-00889​-4 Frör, O., 2008. Bounded rationality in contingent valuation: Empirical evidence using cognitive psychology. Ecological Economics 68, 570–81. https://​doi​.org/​10​.1016/​j​.ecolecon​.2008​.05​.021 Gsottbauer, E., van den Bergh, J.C.J.M., 2014. Environmental policy when pollutive consumption is sensitive to advertising: Norms versus status. Ecological Economics 107, 39–50. https://​doi​.org/​10​.1016/​j​.ecolecon​.2014​.07​.001 Handgraaf, M.J.J., Van Lidth de Jeude, M.A., Appelt, K.C., 2013. Public praise vs. private pay: Effects of rewards on energy conservation in the workplace. Ecological Economics 86, 86–92. https://​doi​.org/​10​.1016/​j​.ecolecon​.2012​ .11​.008 Holmgren, M., Andersson, H., Sörqvist, P., 2018. Averaging bias in environmental impact estimates: Evidence from the negative footprint illusion. Journal of Environmental Psychology 55, 48–52. https://​doi​.org/​10​.1016/​j​.jenvp​.2017​ .12​.005 Kahneman, D., 2003. Maps of bounded rationality: Psychology for behavioral economics. American Economic Review 93, 1449–75. https://​doi​.org/​10​.1257/​000282803322655392 Kahneman, D., Krueger, A.B., Schkade, D., Schwarz, N., Stone, A., 2004. Toward national well-being accounts. American Economic Review 94, 429–34. https://​doi​.org/​10​.1257/​ 0002828041301713 Kaiser, M., Bernauer, M., Sunstein, C.R., Reisch, L.A., 2020. The power of green defaults: The impact of regional variation of opt-out tariffs on green energy demand in Germany. Ecological

Economics 174, 106685. https://​doi​.org/​10​ .1016/​j​.ecolecon​.2020​.106685 Lezak, S.B., Thibodeau, P.H., 2016. Systems thinking and environmental concern. Journal of Environmental Psychology 46, 143–53. https://​ doi​.org/​10​.1016/​j​.jenvp​.2016​.04​.005 Mullainathan, S., Thaler, R.H., 2000. Behavioral economics (Working Paper No. 7948). National Bureau of Economic Research. https://​doi​.org/​ 10​.3386/​w7948 Sagoff, M., 1998. Aggregation and deliberation in valuing environmental public goods: A look beyond contingent pricing. Ecological Economics 24, 213–30. https://​doi​.org/​10​.1016/​ S0921​-8009(97)00144​-4 Schubert, C., 2017. Green nudges: Do they work? Are they ethical? Ecological Economics 132, 329–42. https://​doi​.org/​10​.1016/​j​.ecolecon​ .2016​.11​.009 Tisdell, C., 1990. Economics and the debate about preservation of species, crop varieties and genetic diversity. Ecological Economics 2, 77–90. https://​doi​.org/​10​.1016/​0921​ -8009(90)90014​-L Tomaselli, M.F., Sheppard, S.R.J., Kozak, R., Gifford, R., 2019. What do Canadians think about economic growth, prosperity and the environment? Ecological Economics 161, 41–9. https://​doi​.org/​10​.1016/​j​.ecolecon​.2019​.03​.007 van den Bergh, J.C.J.M., 2009. The GDP paradox. Journal of Economic Psychology 30, 117–35. https://​doi​.org/​10​.1016/​j​.joep​.2008​.12​.001 van den Bergh, J.C.J.M., Ferrer-i-Carbonell, A., Munda, G., 2000. Alternative models of individual behaviour and implications for environmental policy. Ecological Economics 32, 43–61. https://​doi​.org/​10​.1016/​S0921​-8009(99)00088​ -9 Venkatachalam, L., 2008. Behavioral economics for environmental policy. Ecological Economics 67, 640–5. https://​doi​.org/​10​.1016/​ j​.ecolecon​.2008​.01​.018 Victor, P.A., 2010. Ecological economics and economic growth. Annals of the New York Academy of Sciences 1185, 237–45. https://​doi​ .org/​10​.1111/​j​.1749​-6632​.2009​.05284​.x Walstad, W.B., 1997. The effect of economic knowledge on public opinion of economic issues. Journal of Economic Education 28, 195–205. https://​doi​.org/​10​.1080/​00220489709596744

Stefan Drews

8.

Carbon taxes

while that of oxygen (O) is 16, the mass of 1 mole of CO2 is 12 + 2 ∙ 16 = 44 grams. Therefore, one ton of CO2 contains about 12/44 = 0.273 tons of carbon, or conversely, 1 ton of carbon released in the air, once combined with oxygen, corresponds to 3.67 tons of CO2. A carbon tax can thus be formulated as well as a CO2 tax, and the two expressions are in fact often used interchangeably. It is nevertheless crucial to explicitly define whether the tax applies to carbon (C) or to carbon dioxide (CO2), since the rate of a CO2 tax will have to be 3.67 times higher than that of a carbon tax to be equivalent. To make things more tangible, consider the following examples from the two sectors that are among the largest emitters worldwide: transport and housing. When driving a car, burning one liter of gasoline releases 0.64 kg of carbon, which then ultimately results in the emission of 2.3 kg of CO2.2 Hence, driving a car with a fuel consumption of 7 liters/100 km (equivalent to a fuel economy of 34 miles per gallon [mpg]), or the global average fuel consumption of a light-duty vehicle in 2020 (International Energy Agency, 2021), produces 161 grams of CO2 per km, or 1 ton of CO2 every 6200 km. In the residential sector, the average US home consumes about 10 000 kWh of electricity per year,3 roughly the energy delivered by the consumption of 1000 m3 of natural gas. Given that 1 m3 of natural gas contains 0.51 kg of carbon, hence producing 1.9 kg of CO2, heating the average home for a year emits almost 2 tons of CO2. Considering that about 50 percent of households in the United States (Energy Information Administration, 2021) and 40 percent of those in the European Union (Eurostat, 2021) use natural gas for space and water heating, the amount of CO2 produced due to natural gas consumption in the residential sector is considerable. These calculations moreover provide indications of what a carbon tax would represent for fuel consumers. For instance, a tax set at US$80 per ton of CO2 would correspond to about 18 cents per liter (or 70 cents per gallon) of gasoline and about 15 cents per m3 of natural gas.4 These amounts constitute upper bounds for the price increases that may be observed following the implementation of a CO2 tax set at this level.

8.1 Introduction

Carbon taxes, along with emissions trading systems, are instruments for pricing carbon emissions, which account for three-quarters of global greenhouse gas (GHG) emissions, and are therefore the main drivers of climate change. Carbon pricing policies are widely used: to date, 36 carbon taxes and 32 emissions trading systems are implemented worldwide and cover almost 24 percent of global GHG emissions.1 Yet, carbon prices are still low, with only 3.8 percent of global emissions covered by a carbon price above US$40/tCO2, which is the lower bound that may allow the world to meet the 2°C temperature goal of the Paris Agreement (World Bank, 2021). A carbon tax is a surcharge to be paid on a fuel, product, or service in proportion to the quantity of carbon embodied or emitted. By making activities that release carbon in the atmosphere more expensive, the tax creates a financial incentive to limit polluting activities, thereby allowing policy makers to curb emissions and ultimately mitigate climate change. Because they increase the price of certain goods and activities, carbon taxes raise distributional and acceptability concerns. Carbon taxes are thus complex instruments, appealing to chemistry, physics, economics, sociology, and politics. This entry discusses their functioning and main properties.

8.2

Chemical and physical aspects

A carbon tax is a per-unit (or specific) tax, which means that the level of the tax is defined as a fixed monetary amount (USD or another currency) per ton of CO2. Before explaining the economic principles underlying carbon taxes, it is therefore useful to understand what is – concretely – 1 ton of CO2. One ton of CO2 literally weighs 1 ton, or 1000 kg, even though it is a gas. Such a mass of CO2 would fill a cube slightly taller than 8 m (i.e., a volume of about 540 m3). Given that the atomic mass of carbon (C) is 12, 42

Carbon taxes  43

8.3

Economic rationale

From an economic point of view, carbon emissions create so-called “negative externalities” in the sense that agents who participate in the polluting market and make decisions impose costs on others without compensating them. Consider, for instance, the market for fossil motor fuels. Drivers of vehicles with internal combustion engines (ICE) and fuel sellers are the market participants. Pedestrians, bike riders, people whose dwelling is located close to a road, and other non-ICE vehicle owners do not purchase any fossil motor fuel, yet they incur some costs because of fuel consumption by others: they breathe polluted air, which may affect their health. At a broader level, emissions from fossil fuel consumption cause global warming, which impacts everyone, fossil fuel consumers and non-consumers alike. Negative externalities are illustrated in Figure 8.1, which represents the market for a fossil fuel. When the market is unregulated, fuel price depends on the buyers’ willingness to pay and the sellers’ private costs (i.e., the costs incurred by the sellers only). None of the participants consider the external costs, emitting polluting substances in the atmosphere is free, and quantity Q0 fuel will be exchanged at price P0 (point D). The external costs that fall on non-participants in that situation are given by the surface ACDF. The quantity exchanged in the unregulated market is therefore too large, and a deadweight loss equivalent to the surface BCD arises. This situation constitutes a market failure. To achieve economic efficiency, a “Pigouvian tax” (named after Arthur Cecil Pigou, 1920) may be introduced. The rate of the Pigouvian tax corresponds to the marginal external costs. Market participants are hence forced to consider external costs as their own, and they adjust their behavior. The optimal tax rate in Figure 8.1 is given by the distance between point E and point B. Such a tax level would raise the marginal private cost up to the level of the marginal social cost, and the equilibrium of the market would therefore move to point B, where the deadweight loss is completely eliminated. Quantity exchanged would decrease to Q1 whereas market price would increase to P1. Importantly, note that the price increase is usually lower than the tax rate, implying that buyers and sellers share the burden of the tax. Compared to the

free market equilibrium, not only does the price paid by the buyers increase (B is above D), but the price received by the sellers also decreases (E is below D). One may also note that, even in the efficient situation, polluting emissions are not down to zero. The costs for eliminating completely the polluting activities would be larger than the benefits of decreased pollution. When setting the efficient level of a carbon tax, the costs of pollution are to be compared with the benefits of the activities associated with pollution. However, comparing costs and benefits is challenging, in particular in the context of climate change, because damages are difficult to assess and will mostly occur in the future. Several attempts exist in the literature (e.g., Tol, 2019; Taconet et al., 2021), but in general, climate policy objectives are not determined on efficiency grounds. In this context, carbon taxes are instruments that can be used to reach emission reduction objectives cost-effectively (Tietenberg, 1973), or at the lowest global cost.

8.4 Cost-effectiveness

With a carbon tax, property rights on the atmosphere belong to society, and emitters must thus pay a charge for each ton of carbon emitted (or in proportion to the carbon content). Each emitter is then free to decide whether it is worth polluting and paying the tax, or if it is preferable to avoid paying the tax by reducing emissions (i.e., abating). However, abatement activities may be costly because they require investments in more efficient devices, behavioral changes, and so forth. A carbon tax will thus incentivize polluters to compare the costs of emitting and paying the tax versus the abatement costs. To save money, polluters with low abatement costs will have to achieve significant emission reductions. On the contrary, polluters with higher abatement costs will reduce smaller amounts of emissions, but they will conversely pay higher amounts of taxes. This situation leads to cost-effectiveness. More precisely, since the carbon tax rate is unique, each polluter will abate up to the point where the marginal cost of abatement is equal to the tax rate. At higher abatement levels, it would become cheaper to pay the tax. This characteristic of carbon taxes is particularly interesting compared to more traditional approaches, like command-and-control, Andrea Baranzini and Sylvain Weber

44  Elgar encyclopedia of ecological economics

Figure 8.1

Externalities and environmental taxes

in which polluters are assigned reduction targets. Under such a centralized system, the distribution of abatement efforts between polluters will be different from the one obtained with a tax. Most certainly, the total costs for achieving the same abatement level will also be higher since the regulator does not possess precise information on abatement costs and thus cannot allocate abatement efforts consequently. Decentralized pricing mechanisms let emitters make decisions for themselves, thereby making it possible to find cost-effective solutions. Two points must be highlighted. First, with a carbon tax, there is uncertainty regarding the abatement efforts and therefore no guarantee that any policy target will be reached. If the total amount of abatement is lower than the target, the tax rate should be increased. For this reason, modern carbon tax laws (e.g., those in Switzerland) allow tax rates to vary depending on total abatement achieved. As a corollary, if the number of emitters rises, the tax rate must be increased to maintain the policy target. Second, all polluters have to pay the carbon tax on their remaining emissions, even the ones making the greatest abatement efforts. Carbon taxes have, thus, two main consequences: (i) they generate substantial revenues, and how to use these revenues is largely debated in the literature (e.g., Bourgeois et al., 2021); and (ii) they create sustained incentives to abate through time Andrea Baranzini and Sylvain Weber

and may thus contribute to innovation (e.g., Acemoglu et al., 2012). Next to carbon taxes, markets for trading carbon emission allowances constitute another pricing mechanism. In an emission trading system (ETS), a cap is defined for the total amount of allowed emissions and allowances are then distributed to polluters, either free of charge or by auction. Under this approach, polluters must cover each ton of CO2 emitted with the corresponding number of allowances. The demand and supply for allowances will result in an equilibrium price for carbon emissions. As carbon pricing instruments, carbon taxes and ETSs possess a similar rationale and share the main properties (for a survey, see Baranzini et al., 2017). In some circumstances, they may even yield equivalent results. In particular, if the ETS equilibrium price matches the tax level, and if there are no market distortions, emission reductions will be identical with the two instruments, both at the global level and for each individual polluter. Also, for a given amount of abatement, both instruments are cost-effective. The major difference is that, in the case of a tax, the cost of emitting carbon is known in advance, while there is uncertainty regarding the final amounts of abatement and emissions. Conversely, the quantity of pollution is fixed in an ETS – the cap indeed sets the upper bound for emissions – but the cost of carbon emissions will fluctuate according

Carbon taxes  45

to supply and demand of allowances (see Weitzman, 1974, for the seminal comparison of price versus quantity instruments, and Stavins, 2022, for an extensive discussion of the similarities and differences between carbon taxes and ETS).

8.5

Carbon taxes in practice

Carbon taxes may differ along various characteristics that must be defined when a tax is implemented. The first component of a carbon tax is the tax base, which defines which fuel and which activities are subject to taxation. Ideally, 1 ton of carbon should be taxed at the same rate, regardless of the activity generating the emissions, to avoid market distortions and maximize cost-effectiveness. In practice, however, for political, social, or acceptability reasons, some sectors or industries are usually exempted. In Sweden, for instance, where a carbon tax was implemented in 1991, a lower tax rate has historically been applied to firms compared to households. As of 2018, however, the industry rate is the same as the general rate. Since 2008, Switzerland applies a carbon tax covering heating fossil fuels but totally exempting the transport sector. The level of the tax is the second crucial component. A higher tax rate will have a larger impact on the price of polluting activities, thereby inducing larger reductions of emissions. The cost for consumers will, however, increase, which may cause acceptability issues. Carbon taxes are thus often introduced at low levels and then scaled up over time. Sweden and Switzerland, where carbon taxes are currently around US$130/ tCO2,5 launched their taxes at very modest levels and then gradually increased them. The literature provides relatively few empirical evaluations of the effects of carbon taxes on emission reductions. Green (2021) provides a review of the analyses based on real observations in countries or regions that have actually implemented carbon pricing policies. Most studies suggest that aggregate emission reductions from carbon pricing are limited – generally between 0 percent and 2 percent per year. The low level of the tax rate and the presence of exemptions are the common reasons offered as explanations for this low effectiveness. A related concern is that the tax burden will impact households differently. In particular, low-income households are affected

more adversely than others because energy takes a larger share of their budget than for high-income households. Ohlendorf et al. (2020) use a meta-analysis to investigate distributional impacts of carbon taxes and show that most carbon taxes are indeed regressive. The survey by Pizer and Sexton (2019), in which all types of energy taxes are considered, however, finds that this is not always the case and that the regressive impact is generally small. Nevertheless, the “yellow vests movement” that took place in France at the end of 2018, following the announcement of an increase in diesel taxes, illustrates how carbon taxes may matter for their acceptability. The discontent was most intense in rural regions, which often face lower economic development but have to bear a higher fuel tax burden in comparison with large urban centers (e.g., Beck et al., 2016). A further concern is the competition distortion that may hit firms because of carbon tax rate differentials across countries or sectors. Firms located in countries or sectors with relatively high carbon tax rates may indeed face competitive disadvantages compared to firms located in countries or sectors with lower or no taxes, their tax-inclusive prices becoming higher, everything else equal. That may induce local consumers to start buying goods imported from low-tax countries, causing a phenomenon known as “carbon leakage” (e.g., Aichele and Felbermayr, 2015). The magnitude of the competitive disadvantage induced by carbon taxes, however, appears to be small and may even be offset if the tax revenues are recycled to lower employers’ costs for social security contributions (e.g., Andersen and Ekins, 2009). The third component to be defined in the implementation of a carbon tax is the use of the generated fiscal revenues. Revenues can be left unassigned and added to the state’s budget. This approach is relatively unpopular because it implies an increase of a state’s activities, and most people then consider carbon taxes as a pretext to raise fiscal revenues. Alternatively, revenues can be earmarked, that is, assigned to specific objectives. Most countries in fact decide to assign the revenues from carbon taxes to environmental projects. This approach is generally well supported, since people often overlook the incentive impact of carbon taxes and thus expect tax revenues to be earmarked Andrea Baranzini and Sylvain Weber

46  Elgar encyclopedia of ecological economics

to environmental purposes (see Baranzini and Carattini, 2017; Dresner et al., 2006; Kallbekken et al., 2011; Ott et al., 2021). Another possible approach is to use fiscal revenues to decrease other taxes (e.g., on labor or capital), which could give rise to the so-called “double dividend” since reducing distortionary taxes may have positive impacts on economic growth, employment, or technological development in addition to reducing carbon emissions. Freire-González (2018) conducts a comprehensive literature review and concludes that the double dividend remains an ambiguous question that deserves further research. Andrea Baranzini and Sylvain Weber

Notes 1.

See World Bank’s “Carbon Pricing Dashboard” at https://​ca​rbonpricin​gdashboard​.worldbank​.org/​ for updated figures. 2. CO2 emissions coefficients are available from the EIA’s webpage, “How Much Carbon Dioxide Is Produced when Different Fuels are Burned?” (https://​www​.eia​.gov/​tools/​faqs/​faq​.php​?id​=​73​&​t​=​ 11). 3. Obviously, energy used for heating purposes varies substantially according to a number of parameters, such as building size, efficiency of heating system, quality of building’s insulation, or the differential between internal and external temperatures. 4. The upper bound of the price range for 2020 recommended by the World Bank’s High-Level Commission on Carbon Prices Report (2017) is US$80/tCO2. 5. More precisely and in local currencies, the tax level is SEK 1200/tCO2 since 2021 in Sweden and CHF 120/tCO2 since 2022 in Switzerland.

References

Acemoglu D., Aghion P., Bursztyn L., Hemous, D. (2012): The environment and directed technical change. American Economic Review 102: 131–66. Aichele R., Felbermayr G. (2015): Kyoto and carbon leakage: An empirical analysis of the carbon content of bilateral trade. Review of Economics and Statistics 97(1): 104–15. Andersen, M.S., Ekins P. (2009): Carbon-energy taxation: Lessons from Europe. New York: Oxford University Press. Baranzini A., Carattini S. (2017): Effectiveness, earmarking and labeling: Testing the acceptability of carbon taxes with survey data. Environmental Economics and Policy Studies 19: 197–227. Baranzini A., van den Bergh J.C.J.M., Carattini S., Howarth R.B., Padilla E., Roca J. (2017): Carbon pricing in climate policy: Seven reasons, complementary instruments, and polit-

Andrea Baranzini and Sylvain Weber

ical economy considerations. WIREs Climate Change 8: e462. Beck M., Rivers N., Yonezawa H. (2016): A rural myth? Sources and implications of the perceived unfairness of carbon taxes in rural communities. Ecological Economics 124(C): 124–34. Bourgeois C., Giraudet L.-G., Quirion P. (2021): Lump-sum vs. energy-efficiency subsidy recycling of carbon tax revenue in the residential sector: A French assessment. Ecological Economics 184: 107006. Dresner S., Dunne L., Clinch P., Beuermann C. (2006): Social and political responses to ecological tax reform in Europe: An introduction to the special issue. Energy Policy 34(8): 895–904. Energy Information Administration (2021): Natural gas explained – Use of natural gas, U.S. Energy Information Administration. https://​ www​.eia​.gov/​energyexplained/​natural​-gas/​use​ -of​-natural​-gas​.php Eurostat (2021): Energy consumption in households. https://​ec​.europa​.eu/​eurostat/​ statistics​-explained/​index​.php​?title​=​Energy​ _consumption​_in​_households Freire-González J. (2018): Environmental taxation and the double dividend hypothesis in CGE modelling literature: A critical review. Journal of Policy Modeling 40(1): 194–223. Green J.F. (2021): Does carbon pricing reduce emissions? A review of ex-post analyses. Environmental Research Letters 16(4): 043004. International Energy Agency (2021): Fuel consumption of cars and vans, IEA, Paris. https://​ www​.iea​.org/​reports/​fuel​-consumption​-of​-cars​ -and​-vans Kallbekken S., Kroll S., Cherry T.L. (2011): Do you not like Pigou, or do you not understand him? Tax aversion and revenue recycling in the lab. Journal of Environmental Economics and Management 62: 53–64. Ohlendorf N., Jakob M., Minx J., Schröder C., Steckel, J. (2020): Distributional impacts of carbon pricing. Environmental and Resource Economics 78: 1–42. Ott L., Farsi M., Weber S. (2021): Beyond political divides: Analyzing public opinion on carbon taxation in Switzerland. In A. Franzen and S. Mader (eds.), Research handbook on environmental sociology: 313–39. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Pigou A.C. (1920): The economics of welfare. London: Macmillan. Pizer W.A., Sexton S. (2019): The distributional impacts of energy taxes. Review of Environmental Economics and Policy 13(1): 104–23. Stavins R.N. (2022): The relative merits of carbon pricing instruments: Taxes versus trading. Review of Environmental Economics and Policy 16(1): 62–82.

Carbon taxes  47 Taconet N., Guivarch C., Pottier A (2021): Social cost of carbon under stochastic tipping points. Environmental and Resource Economics 78: 709–37. Tietenberg T.H. (1973): Controlling pollution by price and standard system: A general equilibrium analysis. Swedish Journal of Economics 75: 193–203. Tol R.S.J. (2019): A social cost of carbon for (almost) every country. Energy Economics 83:

555–66. Weitzman M. (1974): Prices vs. quantities. Review of Economic Studies 41(4): 447–91. World Bank (2017): Report of the High-Level Commission on Carbon Prices. Washington, D.C.: World Bank. World Bank (2021): State and trends of carbon pricing 2021. Washington, D.C.: World Bank. https://​openknowledge​.worldbank​.org/​handle/​ 10986/​35620

Andrea Baranzini and Sylvain Weber

9. 9.1

Circular economy

and colleagues fall in the Technocentric CE, followed by Reformist Circular Society (12 per cent), Fortress CE (2.5 per cent), and Transformational Circular Society (1.5 per cent). Thus, the most common definitions of the CE term are generally situated within segmented and optimist discourse types. Although several definitions of CE exist (e.g. Kirchherr et al., 2017, identified 114 of them) and the concept is dynamic (i.e. its definition and interpretation are continuously expanding and changing; De Pascale et al., 2021), most definitions consider that the aim of CE is keeping products, components, and materials in use in the economy for as long as possible, while reducing raw material inputs and waste outputs. The European Parliament (2022) defines CE as

The concept

The concept of circular economy (CE) builds on different schools of thought (Ellen MacArthur Foundation, 2013), but its origins are mainly rooted to industrial ecology, and it uses tools from other fields rather than inventing new ones (Bruel et al., 2019; Korhonen et al., 2018). As pointed out by Stahel (2013: 3), “the concept of a circular economy has many similarities with related concepts, such as closed loop economy, lake and loop economy, industrial ecology, cradle to cradle, and material efficiency”. As industrial ecology, CE tries to maximize the value of resources through the economy with the ambition to decouple economic growth and resource consumption (Andersen, 2007; United Nations Environmental Programme [UNEP], 2011). According to Stahel (2016) there are three types of systems thinking in industrial ecology: 1) linear economy (exemplified as a river), 2) CE (exemplified as a lake), and 3) performance economy. The latter would focus on solutions instead of products and would go a step forward from CE by selling services instead of goods through rent, lease, and share business models. CE has been addressed in the literature with different approaches. Calisto et al. (2020) defined four circularity discourse typologies: “Reformist Circular Society”, “Technocentric CE”, “Transformational CE”, and “Fortress CE”. The first two types are considered optimistic with respect to the capacity of technology and innovation to overcome the major ecological challenges of the Anthropocene before an irreversible socio-ecological collapse occurs, where the last two are considered sceptical. In addition, the paper distinguishes between circular society and CE by dividing holistic and segmented discourses. Holistic discourses comprehensively integrate the social, ecological, and political considerations of circularity (the first and third type). Segmented discourses, on the other hand, have a homogeneous perspective and a uniform focus on only economic and technical components of circularity (the 2nd and 4th types). Eighty-four per cent of the 120 definitions revised by Calisto

A production and consumption model which involves reusing, repairing, refurbishing and recycling existing materials and products to keep materials within the economy wherever possible. A circular economy implies that waste will itself become a resource, consequently minimising the actual amount of waste. It is generally opposed to a traditional, linear economic model, which is based on a 'take-make-consume-throw away' pattern.

The first Circular Economy Action Plan package of the European Commission (2015), “Closing the Loop”, stated that CE seeks to maintain the value of products, materials and resources for the longest possible time to develop a sustainable economy that is competitive, less carbon-intensive, and resource-efficient. According to the Organisation for Economic Co-operation and Development (McCarthy et al., 2018), the concept of CE entails: 1) narrowing resource flows to make for more efficient use of natural resources, materials, products, and components in the value chain; 2) slowing resource loops by making more durable products and increasing lifetime through re-use, repair, and remanufacture services; and 3) closing resource loops to minimize raw material extraction and waste output. The French Environment and Energy Management Agency (ADEME) broadens the CE definition to include aspect, such as reducing emissions and improving citizens’ wellbeing, which is not directly mentioned 48

Circular economy  49

in the definitions used by European institutions. ADEME defines CE as an “economic system based around of exchange and production methods that, at every stage of the product life cycle (goods and services), aim to increase the efficiency of resource usage and diminish environmental impact, while also improving the wellbeing of individual citizens” (Monitoring and Statistics Directorate, 2017: 6). Finally, the broadest definition of CE is given by the Ellen MacArthur Foundation (2022), which conceives CE as a “systems solution framework that tackles global challenges like climate change, biodiversity loss, waste, and pollution” that is “based on three principles, driven by design: 1) Eliminate waste and pollution, 2) Circulate products and materials (at their highest value) and 3) Regenerate nature”. As several authors have already pointed out (e.g. Korhonen et al., 2018), fully circular material cycles are not physically possible due to the first and second entropy laws, and benefits from recycling of materials tend to decrease after a tipping point (Andersen, 2007; Ghisellini et al., 2016) since energy input for high-rate recycling can be larger than for raw material extraction. Because of this, it is important to consider the energy flows within the CE definition and implementation. It is also essential to note that recycling alone cannot be seen as a CE solution for the current raw material shortage crisis, and such a concept should be complemented with prevention as well as responsible consumption policies. Although the ambition and expectation included in the CE definitions are high, CE has most often been considered only as an approach to more appropriate waste management (Ghisellini et al., 2016). De Jesus and Mendonça (2018) classify the main barriers to CE found in the literature into: 1) technical limitations (including, for example, inappropriate technology and lag between design and diffusion); 2) economic/ financial/market limitations (e.g. large capital requirements, asymmetric information, and uncertain return and profit); and 3) institutional/regulatory/social/cultural limitations (e.g. misaligned incentives, lacking a conducive legal system, deficient institutional framework, rigidity of consumer behaviour and businesses routines). According to Castro

et al. (2022), the recognized benefits of CE are not always achieved due to the occurrence of rebound effects, which delay the achievement of CE’s full potential.

9.2

Main indicators used for circular economy

To address how the principles of the CE are being implemented at different levels, circularity metrics/indicators are needed. Due to the concept amplitude, a set of indicators rather than a single indicator are needed (Moraga et al., 2019). Such indicators are useful for measuring benefits of CE. Wijkman et al. (2017) estimated that the shift from a linear to a circular economy in five European Union (EU) countries (Finland, France, the Netherlands, Spain, and Sweden) would, for example, make the economy in each country 25 per cent more energy efficient, cut each country’s greenhouse gas emissions by two-thirds or more, and reduce unemployment rates by around one-third in Sweden, the Netherlands, Finland, and France, and by 15–20 per cent in Spain. Efforts around the world are being made to measure progress towards CE. In the EU, the monitoring framework on the CE as set up by the European Commission consists of ten indicators that are divided into four thematic areas (production and consumption, waste management, secondary raw materials, and competitiveness and innovation; Table 9.1). Eurostat indicators are at the Member State level, thus they are considered macro-level indicators for CE, but there are also microand meso-level indicators. In the CE literature, micro-level indicators are used to follow the circularity progress for products, companies, and consumers; meso-level indicators are used for industrial symbiosis systems (e.g. eco-industrial parks); and macro-level indicators for city, province, region, nation, and beyond. According to some authors (e.g. Llorente-González and Vence, 2019), the set of indicators chosen by the European Commission is insufficient and the CE in the EU Monitoring Framework lacks metrics on energy, land, and water use, as well as emissions. In addition, efforts to develop indicators to track the average durability of products, as well as the ecodesign, repair,

Ignasi Puig Ventosa and Verónica Martínez Sánchez

50  Elgar encyclopedia of ecological economics Table 9.1

Indicators used to monitor the circular economy at the EU Member State level

Production and consumption

Waste management

Secondary raw materials

Competitiveness and

Self-sufficiency of raw materials

Recycling rates

Contribution of recycled materials Private investments, jobs and

innovation for production in the EU

to raw materials demand

gross value added

Green public procurement Waste generation

Specific waste streams

Trade of recyclable raw materials Patents related to recycling

Food waste

recycling

between the EU Member States

and secondary raw materials

and with the rest of the world

Source: Eurostat (2022).

re-use, and collaborative consumption activities, are requested by the same authors. Although the growing interest in CE metrics can be observed with the number of publications and the growing number of indicators used (e.g. De Pascale et al., 2021, surveyed 61 CE indicators), standardized methodologies to evaluate CE are still lacking (Kirchherr et al., 2017; Saidani et al., 2019; De Pascale et al., 2021).

9.3

Circular economy and EU regulation

The concept of CE has been progressively introduced into environmental policies worldwide (Kaza et al., 2018; OECD, 2020). This chapter summarizes the situation in this regard in the EU, one of the regions where this has been further developed. In 2015, the European Commission approved its first Circular Economy Action Plan (CEAP; European Commission, 2015), which contained a broad number of actions, including taking actions on Green Public Procurement, proposing improvements to the rules on “end-of-waste”, supporting EU-wide research on raw materials flows, or including guidance on best waste management and resource efficiency practices in industrial sectors through Best Available Techniques reference documents (BREFs). One of the most relevant aspects of the first CEAP was the revision of the main waste directives. Some of the main elements of such revisions were: ● Directive (EU) 2018/850 amending the Landfill Directive (Council Directive 1999/31/EC). By 2035 the amount of municipal waste being landfilled at the national level should be reduced to

a maximum of 10 per cent of the total amount generated. ● Directive (EU) 2018/851 amending the Waste Framework Directive (Directive 2008/98/EC). Inclusion of preparation for re-use and recycling targets for municipal waste by 2020 (50 per cent), 2025 (55 per cent), 2030 (60 per cent), and 2035 (65 per cent); introduction of more concrete requirements in the adoption of Extended Producer Responsibility; harmonization in the way Member States report their results, and so on. ● Directive (EU) 2018/852 amending Directive 94/62/EC on packaging and packaging waste. It updated the recycling targets for different types of packaging waste. ● Directive (EU) 2019/904 of June 5, 2019, on the reduction of the impact of certain plastic products on the environment. It bans several single-use plastic products, such as cotton buds, cutlery, straws, and so on; for other single-use plastics, it established design requirements (recycled content of plastic bottles), marking requirements (e.g. wet wipes or tobacco filters), or separate collection targets (e.g. 77 per cent separate collection of beverage bottles by 2025 and 90 per cent by 2029), and so forth. Some other relevant results from the implementation of the first CEAP were the European Strategy for Plastics in a Circular Economy (European Commission, 2018a), the Report on Critical Raw Materials and the Circular Economy (European Commission, 2018b), and the adoption of ten ecodesign implementing regulations in 2019. The final evaluation of the original CEAP was conducted by the European Commission (2019a).

Ignasi Puig Ventosa and Verónica Martínez Sánchez

Circular economy  51

In March 2020, the European Commission (2020) adopted a new CEAP, which is now one of the main elements of the European Green Deal, approved in 2019 (European Commission, 2019a). The ongoing CEAP includes 35 actions listed in its Annex. Some of the most relevant ones are, for example: development of legislative and non-legislative measures establishing a new “right to repair”, approving an EU Strategy for Textiles, setting waste reduction targets for specific streams, and other measures on waste prevention, revision of the rules on waste shipments, and so on. The new CEAP focuses on seven product value chains, which are considered key from a CE perspective: electronics and information and communication technology, batteries and vehicles, packaging, plastics, textiles, construction and buildings, and food, water, and nutrients. The CEAP acknowledges that, although there have been measures to address certain aspects of the sustainability of products (e.g. the Ecodesign Directive or the EU Ecolabel), “there is no comprehensive set of requirements to ensure that all products placed on the EU market become increasingly sustainable and stand the test of circularity” (European Commission, 2020: 3). In this sense, to ensure that making products aimed at a resource-efficient and CE becomes the norm, in March 2022 the European Commission (2022) launched the Communication “On making sustainable products the norm”, which is a proposal for a regulation on ecodesign for sustainable products aiming at making products more durable, reparable, upgradable, recyclable, and energy and resource efficient.

9.4

Economic environmental instruments to incentivize a more circular economy

Doing things that are bad for the environment cannot come out cheaper. In this sense, to advance towards a CE, different economic signals are needed. This fits within the broader context of the proposals on ecological fiscal reform (e.g. European Environment Agency, 2000, 2022; OECD, 2007). Specifically on the topic of CE, several authors have analysed the role of taxation (e.g. Milios, 2021; Vence and López Pérez, 2021).

Economic policy instruments have the potential to make circularity more economically attractive. A priority should be making more expensive the lower tiers of the waste hierarchy. To that end, landfill and incineration taxes are crucial. These taxes encourage taxpayers (be it local authorities or industrial producers) to adopt strategies to divert waste from landfills and incinerators. These taxes are highly effective (Watkins et al., 2012) and should constitute a central piece of national and regional waste management strategies. Additionally, their revenue can be earmarked and dedicated to supporting preventive waste policies. Several authors suggest that a reform of the VAT could play a significant environmental role in changing consumer preferences (e.g. Copenhagen Economics, 2007; Næss-Schmidt et al., 2008; Oosterhuis et al., 2008; Bahn-Walkowiak and Wilts, 2015; Cannas and Fermeglia, 2021). Specifically, the reform of VAT has also been suggested to favour resource efficiency (RREUSE, 2013) and to support re-use and repair (Köppl et al., 2019; Orón Moratal, 2021). Besides VAT, levying taxes on specific products causing significant environmental impacts could also be a positive measure towards circularity (e.g. the plastic bag tax in Ireland; Anastasio and Nix, 2016). Apart from taxation, another economic instrument is Extended Producer Responsibility (EPR). EPR “aims to make producers responsible for the environmental impacts of their products throughout the product chain, from design to the post-consumer phase” (OECD, 2016). EPR has been most commonly applied to packaging, electric and electronic equipment, batteries and accumulators, end-of-life vehicles, and oils, although this varies from country to country. For most products, and in many countries, EPR does not exist, therefore it should continue to be deployed, especially to products with significant waste generation (e.g. graphic paper, furniture, or textiles) or that are environmentally problematic and costly to manage (e.g. disposable nappies, cleaning wipes, mattresses, cigarette buds, chewing gum, etc). Often, environmental results from EPR have been poor. Eco-modulation of the EPR tariffs is an effective way to incentivize ecodesign among producers. Banning programmed obsolescence and ensuring repara-

Ignasi Puig Ventosa and Verónica Martínez Sánchez

52  Elgar encyclopedia of ecological economics

bility should also be part of the responsibility of the producers. A particular form of materializing EPR is through deposit-refund schemes (DRS). Although they can potentially be applied to other items, their main application so far has been on beverage bottles and cans. These products are sold with a deposit, which is refunded when the empty packaging is returned. This ensures a high level of return (typically around 80–90 per cent), sensibly higher than that achieved in street containers for packaging waste. Around 40 of these schemes are applied in different jurisdictions around the world, mainly in the EU, USA, Canada, and Australia (Zhou et al., 2020). DRS can be applied to both disposable items (to ensure collection and high-quality recycling) and reusables (e.g. glass bottles are cleaned, sanitized, and refilled). Among local authorities, an interesting possibility is transforming waste charges into Pay-As-You-Throw schemes (PAYT), thereby creating an incentive for users towards separate collection (Elia et al., 2015). PAYT requires user identification and measurement of the actual waste generation of each user. Variants of pay-per-bag and pay-per-bin associated with kerbside collection are the most common forms of PAYT; other options involve the use of smart containers (Agència de Residus de Catalunya, 2010). Other economic instruments with significant potential, but which have been less frequently implemented so far, are feebate systems and landfill allowance trading schemes: ● Feebate systems make simultaneous use of fees and rebates. Most municipalities group themselves to manage solid wastes more efficiently, sharing services and facilities. In these associations of municipalities, costs are allocated to each member according to some criteria (e.g. number of inhabitants or amount of waste brought to the shared facilities), which often do not provide sufficient incentives for good practices. In this context, a feebate system could be adequate to reward those municipalities doing better (less per capita generation, higher separate collection rates, higher biowaste quality, etc.), while penalizing the others,

using the average values as a reference (Puig-Ventosa, 2004). ● Landfill Allowance Trading Schemes (LATS) are a form of tradeable instrument used to achieve landfill diversion targets. Allowances for landfilling of municipal solid waste (or the biodegradable part of it) are allocated to local authorities. The quantity of allowances assigned globally is reduced annually to meet the landfill diversion objectives. To achieve their commitments, local authorities can exchange allowances, or may reprofile their own allocation through banking or borrowing. The main experience with LATS was in the UK. Starting in 2005 it had a successful application, but it was abandoned in 2012 when it became redundant with the UK Landfill Tax (Calaf Forn et al., 2014). ● Repair Bonuses subsidize a percentage of repair costs. In Austria, starting in April 2022, the repair bonus subsidized 50 percent of repair costs for electronic and electrical equipment, and it is capped at €200 per repair. This bonus is expected to run until 2026 and to subsidize 400 000 repairs (see https://​www​.reparaturbonus​ .at).

9.5 Conclusions

The concept of CE is not a single and commonly accepted one. It also overlaps with long-established disciplines such as industrial ecology, ecodesign, waste management, material flow analysis, and others. This chapter has presented some of the alternative and sometimes contested views around the concept, and some of the ongoing policies, specifically in the EU. The nature of the policies needed to advance towards a CE should be of different kinds, including communication and awareness, investment in public infrastructures, and of course regulatory measures (e.g. caps on resource extraction). However, due to the economic focus of this book, this chapter has highlighted the role of economic incentives (specifically but not only taxes), which could help align the economy and the environment, and provide economic signals to foster a more resource efficient economic system. Ignasi Puig Ventosa and Verónica Martínez Sánchez

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Circular economy  53

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Ignasi Puig Ventosa and Verónica Martínez Sánchez

10. Climate change and social justice 10.1

quality is running up against the wall of intra-generational inequality here and now: the implementation of a real environmental transition cannot escape the social challenges of the present, especially the imperative of reducing inequalities. This is the great lesson of the “yellow vest” crisis that shook France in 2018–19: the transition will be a just one or will not be. Hence, the two intricate issues that this chapter addresses theoretically but also assesses empirically are: climate justice understood as global social justice (or justice between) and climate justice understood as national social justice (or justice within). These theoretical and empirical considerations first need to be embedded in a broader analytical framework linking climate change and social justice.

Climate crisis as social injustice

All meaningful conversations among humans about reform and progress begin with a robust debate about the principles of justice people want to see upheld and the institutions capable of implementing them. This is especially true of the titanic change in attitudes and behaviors required to avoid climate disaster. Why is the climate crisis worsening before our very eyes, even though we have all the scientific, technological, and economic tools we need to extricate ourselves from fossil fuels? Largely because those most responsible are not the most vulnerable, and vice versa. On the one hand, a handful of countries – around 10 percent (and a small minority of people and industries within these countries1) – are responsible for close to 80 percent of greenhouse gas (GHG) emissions, causing climate change that is increasingly destroying the well-being of a considerable part of humanity across the world, especially in poor and developing countries. On the other hand, the vast majority of people most affected by climate change (in Africa and Asia), numbering in the billions, live in countries that bear almost no responsibility for it. Yet these same countries are extremely vulnerable to the disastrous consequences of climate change (heat waves, hurricanes, and floods) caused by the lifestyles of other people living thousands of miles away. The entire African continent accounts for around 3 percent of global GHG emissions, but its population will be massively exposed to water shortages attributable to climate change in coming decades. Climate justice is therefore key to understanding and eventually solving the climate crisis. What is true in space, between countries, is also true in time, between generations. Thanks to Greta Thunberg and the youth movement she started, climate strikes and protests are gaining traction and having impact on public debates. Some people among the new generations are now aware of the grave injustice they will suffer as a result of choices over which they do not yet have any control. But recognition of this intergenerational ine-

10.2

Social justice and climate change: an analytical framework2

The Intergovernmental Panel on Climate Change (IPCC; Pörtner, 2022: 7) defines social justice and climate justice in the following way: “Social justice comprises just or fair relations within society that seek to address the distribution of wealth, access to resources, opportunity and support according to principles of justice and fairness. Climate justice comprises justice that links development and human rights to achieve a rights-based approach to addressing climate change.” Of course, many conceptions of justice co-exist and determine different streams of environmental justice able to provide a robust basis for such a rights-based approach. One of them consists in embracing the capability-building and human development framework developed by Amartya Sen. In essence, the capability approach recommends that well-being be assessed beyond material conditions and also reflect the quality of life of a given person. Based on Sen’s analytical framework, one can define an environmental inequality as a situation that results in an injustice, or that is unjust if the well-being and capabilities of a particular population are disproportionately affected by its environmental conditions of existence (Laurent, 2021a). The environmental conditions of existence consist of, negatively, exposure to pollution and risks, and, positively, access 56

Climate change and social justice  57 Table 10.1 A typology of climate inequality  Types of air Philosophical

Generative fact

inequality

approach

Type 1

Procedural justice Impact of individuals

Inequality vectors

Inequality criteria

Exclusion from public

 

Climate inequality examples

and groups on climate

Non-participation

decision-making procedures

in climate adaptation policy

policies Type 2

of localities Vertical and

Recognitive

Impact of climate

Taxation, regulatory

Nationality, spatial

justice*

policies on individuals

policies, information/

location, age, gender, horizontal income

and groups

awareness

socio-economic level inequalities caused by carbon taxation

(income, health, education, etc.), ethnic characteristics, and so on Type 3

 

 

Unequal exposure

Distributive

Exposure/sensitivity

justice

(vulnerability) to climate

and sensitivity to

change

heatwaves in urban areas

Type 4

Distributive

Impact of individuals

CO2, health related climate

justice

and groups on climate

impacts, and so on

change

 

Carbon footprint by the top income deciles

Source: Adapted from Laurent (2022). Note: *This is a process model of social justice that includes a positive regard for social difference and the centrality of socially democratic processes. For a large number of examples of climate inequalities, see Guivarch and Taconet (2021).

to amenities and natural resources (water, air, food). The particular character of the population in question can be defined according to different criteria: social, demographic, territorial, and so on. On this basis, different categories of climate inequality exist and must be broken down to be properly identified and possibly addressed and mitigated. A first typology of climate inequalities regarding their generative factor (the event generating the inequality) consists in dividing them into two categories: the inequality impact of individuals and groups on climate damage and definition of climate policies, and the inequality impact on individuals and groups, by climate policies and climate damage. A second typology of climate inequalities consists in considering their inequality vector: what form of climate degradation is responsible for the observed injustice. A third typology looks at criteria of inequality: what dimension of human beings is at play in the observed injustice. Table 10.1

summarizes this framework and highlights four types of climate inequality: ● Type 1 is concerned with procedural justice and stems from the potential exclusion of individuals and groups from public policy procedures, for instance, the inability to participate in climate adaptation strategy; ● Type 2 is concerned with recognitive justice and stems from potentially adverse social effects of climate change mitigation and adaptation policy on individuals and groups, for instance, the regressive impact of energy taxation on poorer households; ● Type 3 is concerned with distributive justice and stems from the unequal exposure and sensitivity of individuals and groups to climate change impact, for instance, the heavier burden placed on disadvantaged isolated groups in metropolitan areas during heatwaves; and ● Type 4 is also concerned with distributional justice but stems from the unequal responsibility of individuals and groups in climate change, for instance, the greater carbon footprint of richer households. Éloi Laurent

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10.3

Climate justice as global social justice

Global climate justice has numerous and complex meanings in the existing academic literature (Bourban, 2021). Here, I limit myself to one essential question: Who has the right to consume the remaining carbon budget until 2050 and on what basis? The nature of “who” is determined by the way climate negotiations work: because the countries that are parties to the United Nations Framework Convention on Climate Change negotiate climate targets and efforts, any realistic allocation framework should end up determining national targets. In light of the IPCC’s Special Report on 1.5° published in 2018 (Masson-Delmotte et al., 2018), it is possible to determine the global carbon budget: in 2019, it amounted to 945 GtCO2e, corresponding to an intermediate target between 1.5° and 2° associated with the 67th percentile of the Transient Climate Response to Emissions, in line with the goals set in Article 2 of the Paris Agreement. The question of the fair distribution of this global carbon budget has been the subject of numerous studies (for a summary and proposals, again see Bourban, 2021), but there is currently no work that integrates a complete vision of the three justice criteria identified in the academic literature – equity, responsibility, and capacity (Höhne et al., 2014) –to determine an operational distribution of national efforts or so-called “effort sharing.” I thus focus on defining operational indicators for all three of these climate justice dimensions, paying attention to their logical order, statistical robustness, and straightforwardness. My framework can be summarized in a simple step-by-step justice procedure starting with the biophysical constraint of global carbon budget and allocating the resulting universal carbon endowments through simple criteria using equity, responsibility, and capacity indicators. Each country receives an initial carbon endowment that is modulated (adjusted) using first equity, then responsibility, and finally capacity. With this framework in mind, I focus on the top 20 emitting countries, which accounted for 77 percent of emissions in 2019.3 I assume that the emissions reduction target will be shared by all countries by 2050 and that the carbon budget therefore covers Éloi Laurent

the next 30 years, which translates into an average annual budget of around 30 GtCO2e (for comparison, 36 GtCO2e were emitted in 2019). I take as a starting point an equal distribution among all members of humanity in 2019, meaning an initial allocation of 122.5 tCO2e up to 2050 (i.e., about 4 tCO2e per person-year – a country’s budget being the aggregation of the individual allocations of its total population). I interpret the equity criterion as meaning that the world’s citizens all have equal access to the GHG storage capacity of the atmosphere (this corresponds to a universal carbon endowment corrected for each major emitter for its population and for population growth until 2050). The responsibility criterion is the amount of GHGs already emitted since 1990 in consumption, thus combining a spatial justice criterion with a temporal criterion, reflecting the global as well as the historical responsibility of individual countries. Finally, the capacity criterion is expressed here by the United Nations Human Development Index (HDI), which by construction ranges from 0 to 1, and which is related for each country to the world average (which in 2019 was 0.737). Countries whose HDI is lower than the world average see their budget increase in proportion to their human underdevelopment, and vice versa for developed countries. Applying these three criteria leads to allocating a just carbon budget per and per capita for each of the top 20 emitters (Figure 10.1). Looking at Figure 10.1, one can see that the equity criterion generally operates a reallocation from countries with a falling population to those with a rising population, which are almost entirely located in sub-Saharan Africa. In this respect, based on this criterion, China undergoes a reduction in its budget of 44 GtCO2e (almost 25 percent), while the rest of the world benefits from an increase of 86 GtCO2e. The responsibility criterion appears to be the main determinant leading to a reallocation of the global budget between countries, with a transfer of nearly 263 GtCO2e from the Organisation for Economic Co-operation and Development countries to the so-called developing countries. The capacity criterion also leads to a reallocation toward developing countries, but much less (almost 34 GtCO2e in total). Thus, each criterion plays out differently (either by the nature of the rebalancing or by its extent), suggesting that the interplay of

Climate change and social justice  59

Source:

UN Global Carbon Project, author’s calculations.

Figure 10.1 Global carbon budget allocation using three justice criteria

this relatively simple set of three criteria does indeed enable different understandings or conceptions of climate justice. This aggregate allocation can be translated into a distribution of the burden of the mitigation effort for all 20 top emitters (Figure 10.2). In light of these results, it is clear that developed countries have a climate debt in the form of negative emissions, meaning that they must not only cut emissions to zero but then help cut emissions in accordance to their remaining carbon debt in countries that have a positive remaining carbon budget (for a recent literature review and different empirical outcomes regarding justice considerations in carbon budget allocation, see Williges et al., 2022). The same justice lens could be and in fact has been applied to adaptation financing (for a review, see Khan et al., 2020). However, this first step toward climate justice does not tell anything about the intersection of social justice and climate change within each country: each country should thus extend the logic of climate justice within its national frontiers in order to allocate globally determined national carbon budgets to social groups and down to individuals.

10.4

Climate justice as national social justice

The question I now turn to is indeed the national climate justice strategy: On what basis can a given country allocate its globally determined national carbon budget? According to Figure 10.2, France has to cut 2.3 tons per capita per year until 2050; how can this be done in an equitable way according to France’s own justice principles and mitigation policies? Taking the carbon distribution as computed by Malliet (2020) based on the French Household Budget Survey data collected by Institut national de la statistique et des études économiques (INSEE) in 2011, one can determine how the aggregated carbon footprint is shared among the French population with respect to income distribution carbon footprint inequality translating by definition into larger reduction efforts being demanded from the top four deciles in the perspective of global climate justice developed in the first section (Figure 10.3). The magnitude of the effort expected to comply with biospheric limits and global justice principles is such that these reductions in emissions can appear unrealistic Éloi Laurent

60  Elgar encyclopedia of ecological economics

Note: Each bar indicates the effect of each criterion, taken independently of the others, on the average annual carbon budget per country. For example, while each American citizen has an initial allocation of 4 tCO2e, the equity criterion leads to this budget being reduced to 3.73 tCO2e, the application of the responsibility principle leads to the initial allocation turning negative and corresponding to a debt of 13 tCO2e, and the capacity criterion reduces the initial allocation to 3.25 tCO2e. The aggregation of these different criteria results in a total negative budget of 9.5 tCO2e per capita per year. A negative budget here reflects the fact that the historical emissions taken into account via the responsibility criterion are higher than the current carbon budget allocated via the other criteria. Source: UN Global Carbon Project, author’s calculations.

Figure 10.2 National just carbon budgets, in TCO2e per capita and per year

and unfeasible. In fact, the national carbon strategy (Stratégie nationale bas carbone, or SNBC) adopted in its latest version in 2019 recommends annual cuts on the order of 5 percent, while Figure 10.3b implies annual cuts close to 20 percent. But the SNBC does not account for biospheric limits, or historical responsibility, or global impact of French carbon consumption (it is based on national/production emissions cuts in line with a general European Union target loosely based on climate science). What is more, the considerable effort of getting to zero emissions can be spread over a longer period of time, resulting in smaller annual cuts, with a financial compensation matching this extension in the period post-zero, where France would commit to cutting emissions in countries with positive carbon budgets. Even more importantly, to be sustainable, this effort should be calibrated among differÉloi Laurent

ent social groups based on national climate justice principles. The mitigation effort is by definition steeper for higher income deciles, but this is clearly not fair enough. A possible just transition strategy could imply designing a tax and transfer policy able to connect reductions in “luxury emissions” (air and road leisure travel, luxury consumption4) to reductions in “essential emissions” (food, housing, and work mobility). Important cuts can indeed be achieved quickly in the carbon footprint of French social groups. A round-trip flight to New York is worth a ton of CO2e, the same as six round-trip flights from Paris to the South of France (luxury emissions), which is equivalent to a year of home heating or the average emissions of a car used to drive 5000 km (essential emissions).5 Using equity and capacity as guiding justice principles (leaving out the shared

Climate change and social justice  61

Source:

Malliet (2020) and author’s calculations.

Figure 10.3 Annual emissions reduction by income decile in France, in tCO2e per capita, on average

collective historical national responsibility of the French population), one could imagine a social-ecological progressive taxation system whereby the carbon footprint of the higher deciles would be heavily taxed and reduced, generating important revenues to finance the reduction in emissions for lower deciles via public investment (for instance, in home retrofit). Designing these social-ecological progressive tax policies using income and location as justice criteria is clearly feasible in France and beyond (see Berry and Laurent, 2019, for France, and Andersson and Atkinson, 2020, for other countries). As is well known, introducing social compensation based on income level but also location (rural areas versus urban areas, suburban areas vs. urban centers, etc.) can maintain the environmental efficiency of the policy measure (compensation should not be understood as exoneration) while easing and even erasing its social regressive impact and therefore increasing its political acceptability as well as fairness (see Laurent, 2011). On the contrary, introducing carbon taxation without social compensation is likely to trigger political opposition and even social protest. In France, the revolt of the so-called

“yellow vests” that shook the country in the fall of 2018 and early 2019 precisely started because a protest against a rise of fuel prices evolved into a social-ecological revolt against the unfair social effect of a planned rise in carbon taxation, taking place against the backdrop of widespread fuel poverty, an environmental inequality which public policy has not recognized until recently and is still unable to curb. Close to 5.5 million French households (i.e., 8 million people) are currently estimated to suffer from fuel poverty (close to 15 percent of the French population), with over 40 percent of households of the first income quartile considered fuel poor. Consequently, the increase of the French carbon tax was clearly socially unjust in three measurable ways: it created vertical inequality, horizontal inequality, and finally increased fuel poverty stemming from the tax and pre-existing inequality (Berry and Laurent, 2019). But all of these inequities can be mitigated and even reversed. Appropriate social compensation appears to be both minimal in cost and easy to implement. Many countries and localities (such as the Nordic countries, but also Indonesia) have indeed successfully introduced such compensations, for instance, Éloi Laurent

62  Elgar encyclopedia of ecological economics

the province of British Columbia, where a carbon tax was rejected by 43 percent of its residents when it was introduced without social compensations in 2008 and is now supported by a large majority (support grew when compensations were introduced). For France, many options of progressive social-ecological taxation exist. For instance, increasing the currently frozen carbon tax to 55 euros per ton of carbon in 2021 as an environmental objective and redistributing 25 percent of revenues to households using existing mechanisms, a majority of households (more than 50 percent of households in the first six deciles of standard of living) could gain from carbon taxation (receiving more in social transfers than what they pay in carbon taxation). The 75 percent of the remaining revenue could be allocated to mitigating fuel poverty but also to provide financial help to shift to low-carbon equipment, reducing social inequality further in a context of rising energy prices (Berry and Laurent, 2019). Hence, progressive social-ecological tax policies may be able to both lower the carbon footprint of the highest deciles while redistributing money to compensate the reduction of lower deciles and allowing them to invest in low-carbon lifestyles. More generally, mitigation and adaptation climate policies should be aligned with the notion of “just transition” (Bauler et al., 2021). Such a transition should no longer be understood only as social support or financial compensation for workers of fossil fuel industries, but more broadly as an integrated social-ecological transition strategy focused on three priorities: 1. Systematically analyze ecological shocks (such as heatwaves, flooding, zoonosis, etc.) and the policies intended to mitigate them from the angle of social justice in its three fundamental dimensions: recognition, distributive, and procedural; 2. Give priority in the design of these just transition policies to dynamic human well-being over economic growth (see Laurent, 2021b); and 3. Build and implement these just transition policies in a democratic way by ensuring the understanding, support, and active participation of citizens.

10.5

On page 18 of the Summary for Policymakers of the Working Group I contribution to the Sixth Assessment Report by the IPCC (Masson-Delmotte et al., 2021), the second column shows that all of the five main climate scenarios considered converge toward a 1.5°C world at a more or less rapid pace. In the same table, the third line shows that one climate scenario, dubbed “SSP1-1.9” (SSP standing for “Shared Socioeconomic Pathways”) foresees a stabilization of global warming at 1.6°C between 2041 and 2060 before witnessing a decrease to 1.4°C at the end of the 21st century. Riahi et al. (2017) have defined the SSP1 scenario in the following terms: Sustainability – Taking the Green Road (Low challenges to mitigation and adaptation) The world shifts gradually, but pervasively, toward a more sustainable path, emphasizing more inclusive development that respects perceived environmental boundaries. Management of the global commons slowly improves, educational and health investments accelerate the demographic transition, and the emphasis on economic growth shifts toward a broader emphasis on human well-being. Driven by an increasing commitment to achieving development goals, inequality is reduced both across and within countries. Consumption is oriented toward low material growth and lower resource and energy intensity. (157, emphasis added)

If SSP1 is correctly understood as a possible path away from climate chaos, then it clearly translates into two important challenges: prioritizing well-being instead of gross domestic product growth (again, see Laurent, 2021b) and reducing inequality both between and within countries. Éloi Laurent

Notes 1.

2. 3.

Éloi Laurent

Conclusion: climate justice as a way out of climate chaos

According to the 2020 estimate by Oxfam, From 1990 to 2015, the richest 10 percent of the world’s population (about 630 million people) were responsible for 52 percent of the cumulative carbon emissions, while the bottom 50 percent (about 3 billion people) were responsible for just 7 percent of cumulative emissions (see Gore, 2020). This part is based on Laurent (2022). The top 20 emitting countries in 2019 were: the United States, Canada, Saudi Arabia, Australia,

Climate change and social justice  63 adaptation finance through a climate justice lens, Climatic Change 161, 251–69. Laurent É., 2011. Issues in environmental justice within the European Union, Ecological Economics, 70(11), 1846–53. Laurent É., 2021a. From the welfare state to the social-ecological state, in É. Laurent and K. Zwickl (eds), The Routledge Handbook of the Political Economy of the Environment, 211–22. Routledge. Laurent É., 2021b. From welfare to farewell: the European social-ecological state beyond economic growth, ETUI Research Paper - Working References Paper 2021.04. http://​dx​.doi​.org/​10​.2139/​ssrn​ Andersson J., and Atkinson G., 2020. The dis.3873766 tributional effects of a carbon tax: The role Laurent É., 2022. Air (ine)quality in the European of income inequality. Centre for Climate Union, Current Environmental Health Reports Change Economics and Policy Working Paper 9(2), 123–9. https://​doi​.org/​10​.1007/​s40572​ 378/Grantham Research Institute on Climate -022​-00348​-6 Change and the Environment Working Paper Malliet P., 2020. L’empreinte carbone des ménages 349. London School of Economics and Political français et les effets redistributifs d’une fiscalité Science. carbone aux frontières [The carbon footprint of Bauler T., Calay V., Fransolet A., Joseph M., French households and the redistributive effects Laurent É., Reginster I., 2021. La transition of carbon taxation at borders], Policy Brief 62. juste en Europe: mesurer pour évoluer [The OFCE. just transition in Europe: measuring and evolv- Masson-Delmotte V., Zhai P., Pirani S.L., ing]. Iweps. https://​www​.iweps​.be/​publication/​ Connors C., Péan S., Berger N. . . . Zhou B. la​-transition​-juste​-en​-europe​-mesurer​-pour​ (eds), 2021. Climate change 2021: The phys-evoluer/​ ical science basis. Contribution of Working Berry A., and Laurent É., 2019. Taxe carbone, Group I to the Sixth Assessment Report of the le retour, àà quelles conditions? [The Carbon Intergovernmental Panel on Climate Change. tax come back, at what conditions?]. Working Cambridge University Press. paper June, 2019, Sciences Po, Observatoire Masson-Delmotte V., Zhai P., Pörtner H.-O., Français des Conjonctures Economiques Roberts D., Skea J., Shukla P.R., . . . Waterfield (OFCE). https://​www​.ofce​.sciences​-po​.fr/​pdf/​ T. (eds), 2018. Global Warming of 1.5°C.An dtravail/​OFCEWP2019​-06​.pdf IPCC special report on the impacts of global Bourban M., 2021. Promoting justice in global warming of 1.5°C above pre-industrial levels climate policies, in É. Laurent and K. Zwickl and related global greenhouse gas emission (eds), Routledge Handbook of the Political pathways, in the context of strengthening the Economy of the Environment, 226–42. global response to the threat of climate change, Routledge. sustainable development, and efforts to eradGore T., 2020. Confronting carbon inequalicate poverty. Intergovernmental Panel on ity - Putting climate justice at the heart of Climate Change. the COVID-19 recovery. Oxfam. https://​ Oswald, Y., Owen, A., Steinberger, J. K. 2020. oxfamilibrary​.openrepository​.com/​bitstream/​ Large inequality in international and intranahandle/​10546/​621052/​mb​-confronting​-carbon​ tional energy footprints between income groups -inequality​-210920​-en​.pdf and across consumption categories, Nature Guivarch C., and Taconet N., 2021. Global ineEnergy, 5(3), 231–239. https://​doi​.org/​10​.1038/​ qualities and climate change, in É. Laurent s41560​-020​-0579​-8 and K. Zwickl (eds), The Routledge Handbook Pörtner H.-O., Roberts D.C., Tignor M., of the Political Economy of the Environment, Poloczanska E.S., Mintenbeck K., Alegría A., 90–103. Routledge. . . . Rama B. (eds), 2022. Climate change Höhne N., den Elzen M., and Escalante D., 2014. 2022: Impacts, adaptation, and vulnerability. Regional GHG reduction targets based on Contribution of Working Group II to the Sixth effort sharing: a comparison of studies, Climate Assessment Report of the Intergovernmental Policy 14, 122–47. Panel on Climate Change. Cambridge Khan M., Robinson S., Weikmans R., Clipet D., University Press. and Roberts J.T., 2020. Twenty-five years of Riahi K., van Vuuren D.P., Kriegler E., Edmonds J., O’Neill B.C., Fujimori S., Bauer N. et al., 2017. The shared socioeconomic pathways and their energy, land use, and greenhouse gas

Germany, Japan, Russia, the United Kingdom, Italy, South Korea, Poland, France, South Africa, Iran, China, Mexico, Turkey, Brazil, Indonesia, and India. I also include the 27-member European Union to provide a basis for comparison. These top-20 emitters represent, on average, 57 percent of the world’s population until 2050 (61 percent in 2020 and 53 percent in 2050). 4. For an empirical study of energy consumption inequality, see Oswald et al. (2020) 5. See DGAC (Direction générale de l’aviation civile).

Éloi Laurent

64  Elgar encyclopedia of ecological economics emissions implications: An overview, Global Environmental Change 42, 153–68. Williges K., Meyer L.H., Steininger K.W., and Kirchengast G., 2022. Fairness critically con-

Éloi Laurent

ditions the carbon budget allocation across countries, Global Environmental Change 74, 102481.

11. Coevolution (socio-biophysical coevolution)

range from highly materialistic, individualistic, and utilitarian to highly other-oriented, communal, and caring, with these different types of value-orientation being selected upon by the dominance of types within the other systems. The individualistic emphasis of utilitarian approaches to thinking about values selects for markets within social organization. Coevolution in ecological economics is a broad way of thinking that relies on and provides a way of connecting knowledge from ethics, organizational theory, epistemology and the history of science, and technology studies, for example. Coevolutionary thinking is a different and additional way of comprehending the interactive history of people and nature. It complements other ways of thinking by providing new insights, while its insights can also contradict other ways of thinking about people and nature. To acknowledge multiple ways of framing reality is to acknowledge the possibilities that the insights from the different framings will be contradictory. For example, while the mechanistic framework of markets helps us see the importance of using resources and managing our environmental interactions efficiently, the coevolutionary framework helps us see the importance of maintaining diversity in both natural and social systems. Diversity, however, comes at the cost of efficiency, as seen by economists. Coevolution also provides a new critical and constructive way of thinking about the rise of the market economy as a dominant form of social organization as well as how democratic social organization transformed into corporatocracy. Socio-biophysical coevolutionary frameworks are historical and typically include social organization beyond markets. For this reason, the coevolutionary framework especially complements historical and institutional approaches to thinking about people and nature. With different ways of knowing incorporated in the knowledge system, and with diversity being important to further coevolution, the socio-biophysical coevolutionary framework provides its own rationale for methodological pluralism (Norgaard, 1989; Goddard et al., 2019). While methodological pluralism has been central to the field of ecological economics, it has also presented problems for defining and mobilizing the field (Spash, 2013). In addition, and con-

Coevolution provides a dynamic systems framework that Richard Norgaard (1981, 1994) and John Gowdy (1994) adapted to address the social and biophysical questions raised in ecological economics. Species evolve to better fit their environment, and other species are a critical part of every species’ environment. Thus, species evolve in response to each other’s evolution (i.e., they coevolve; Ehrlich and Raven, 1964). Charles J. Lumsden and Edward O. Wilson (1981) sought to resolve the socio-biology debate (Jumonville, 2002) by arguing that people and cultures coevolved. Robert Boyd and Peter Richerson (1988) and William Durham (1991) provided a coevolutionary theory of cultural change and human diversity. Coevolution in ecological economics is closely linked with that in anthropology and thereby expands the questions raised in ecological economics (Gowdy, 2021). Norgaard framed social-biophysical coevolution as consisting of five systems: the organizational, knowledge, value, technological, and environmental systems, with the characteristics of each coevolving in response to changes in the characteristics of the others, as suggested in Figure 11.1. Note that the arrows represent both direct interactions and selective evolutionary processes. Thinking in a socio-biophysical1 coevolutionary framework facilitates heuristic explorations of how complex systems directly interact in ways comparable, for example, to predator–prey dynamics in ecology, while they also evolve in response to changes in the characteristics of each other. Interdependence can entail strict reciprocity (mutual causal influence), diffuse reciprocity (where multiple feedbacks simultaneously take place at different levels in each system), or partial reciprocity (only one system causally reacts; Gual and Norgaard, 2010). Within social organization, for example, imagine that its types include markets, forms of governance, modes of community, and models of the family. Each of these types of social organization are being selected on by the dominance of social organizational types, such as markets, for example. Similarly, types of value-orientation can 65

66  Elgar encyclopedia of ecological economics

Source:

Norgaard (1994).

Figure 11.1 The coevolutionary pentagram

troversial in a different way, the framework proposed by Norgaard (1994) has its own internal epistemology that puts the coevolutionary histories of Indigenous and experiential knowledges and of modern science on comparable epistemological footing. The socio-biophysical coevolutionary framework helps us see how values, social organization, technology, and nature all affect how societies and specialized groups within them understand morality and reality and work within these understandings. The atomism of modern science complements the disciplinary organization of science and corresponding management agencies. In parallel, it also helps explain the rise and adoption of discipline-rooted technologies that keep having “surprising” social and environmental systems consequences. In short, as an alternative framework, coevolutionary ecological economics provides insights into how atomistic-mechanistic frameworks, with their seeming ability to predict, so often lead to unexpected and sometimes disastrous Miquel A. Gual and Richard B. Norgaard

results leading to the need for whole systems rethinking (Haider et al., 2021). Evolutionary thinking in the social sciences has been fraught with peril, from racist eugenics and social Darwinism to, for example, arguments that individualism and selfishness are natural because that is how people have coevolved to behave. The perils of past usage of evolution in the social sciences have been closely tied to ideas of progress and superiority intertwined with the phrase “survival of the fittest.” These have led to arguments that some people and social systems have evolved (i.e., progressed) to be more fit and thereby superior to others and will be forever more. Fitness in coevolutionary theory, however, is a product of history that keeps on changing without direction. While cockroaches have proved fit for millennia, most species have coevolved to extinction. Coevolution is about change; there is no inherent progress over time. Coevolution in ecological economics has stressed how present conditions have coevolved and avoided justifying present people, cultures, and institutions. Given

Coevolution (socio-biophysical coevolution)  67

humanity’s current socially and environmentally destructive course, the whole idea of progress has been laid waste. Socio-biophysical coevolution has its own complexities when looking at (mostly) recursive interactions among its dimensions. Gual and Norgaard (2010) elaborated three broad coevolutionary modes: systemic, selective, and adaptive coevolution. Systemic coevolution emphasizes how the characteristics of the social system and their evolution impact the biophysical system, changing biophysical dynamics. In turn, these changed biophysical characteristics affect how selective forces operate, potentially giving way to adaptation and differential evolution of one or several species or populations in affected geographic ranges. Finally, the social system evolved responding (or not) to this evolutionary change, doing it at different scales and system levels. Selective coevolution addresses how specific technologies emerging from the social evolutionary process directly acted as selection forces in the biophysical evolutionary process, mostly accelerating phenotypic and genotypic adaptation. After which, the social system technologies could adapt to this change through cumulative learning. This mode of coevolutionary explanation has been recently used and expanded to explain the coevolutionary process behind antibiotic and pesticide resistance (Søgaard-Jørgensen et al., 2020), and governance responses (Søgaard-Jørgensen et al., 2019). Finally, the mode of forced coevolution through genetic manipulation (or biotechnological adaptive coevolution) referred to the evolutionary process by which the social system develops biotechnologies capable of surpassing natural selection and directly act as adaptation mechanisms in the evolution of biological organisms, with no real knowledge about the reciprocal adaptive futures or possibilities. An ecosystem is a mental construct with arbitrary areal and temporal boundaries. Because the ecological system’s dynamics occur simultaneously at different scales, three great challenges for our understanding arise: the way processes change across scales, the influence of our own perspective on the observed dynamics, and the arising inter-relations of processes among scales (Levin, 2005). Social systems too are mental constructs. The framework in Figure 11.1, for example,

would be appropriate for a society on an island, but societies around the globe have been well connected for centuries. The first step is to acknowledge that there are introductions from one society to another, and power-induced invasions must also be considered. The next step after realizing that there are many coevolving societies is to envision hierarchies of coevolving socio-biophysical systems operating across geographies with both mechanistic and evolutionary processes occurring on different time scales (Gual and Norgaard, 2010).

Applications

To date, coevolutionary ecological economic thinking has been mostly heuristic. Norgaard (1984) argued that the development of Europe and North America was a coevolutionary process and that Amazon development planners, rather than seeking a best plan, would be more successful if they set up diverse development experiments, comparable to genetic mutations or introductions in biological systems, to enhance the possibilities that one experiment would prove fit. Agricultural practices and their governance have proven especially amenable to coevolutionary interpretation (Norgaard, 1981; Saifi and Drake, 2008; Moreno-Peñaranda and Kallis, 2010; Ríos-Núñez et al., 2013; Shiki et al., 2013; Søgaard-Jørgensen et al., 2020). Coevolution has also been used to frame the transition to a renewable energy future (Gual, 2005; Foxon, 2011). More recently, work about socio-biophysical coevolution has come from overlapping and adjacent fields to ecological economics. These range from cultural evolution theory and the Cultural Multilevel Selection framework (Brooks et al., 2018), to coupled coevolutionary and resilience studies (Haider et al., 2021), to Urban and Niche Ecology, with many empirical applications of the eco-evolutionary dynamics framework (Alberti et al., 2017; Miles et al., 2021). All of them show the renewed interest in and pertinence of the process of socio-biophysical coevolution for improving our understanding of how our social system drives and, in many ways, determines alternative futures of systemic reciprocal influence, like climate change, or of systemic, selective, or adaptive (maybe reciprocal) influence on biological diversity. Finally, Gowdy (2021) and Norgaard (2021) argue from a perspective Miquel A. Gual and Richard B. Norgaard

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of human evolution that the broad public acceptance of the assumptions of economics account for the modern destruction of people and planet. We may summarize that the term coevolution is a non-excluding, overarching, cosmological, and epistemological concept useful to explore, among others ways of knowing, recursive interactions among coupled complex systems. As a heuristic, it allows us to identify and focus on specific key elements of the unfolding socio-biophysical coevolutionary change. Finally, we propose an expanded definition of socio-biophysical coevolution as the process of interdependent change taking place at multiple scales and (hierarchical) levels among interrelated entities within complex evolutionary social systems and complex evolutionary biophysical systems. Its actual and future study requires both trans-disciplinarity and consilience within the limits of methodological pluralism. Miquel A. Gual and Richard B. Norgaard

Note

1. The term “socio-biophysical” can be found in the literature, as well as “socio-environmental” or “socio- ecological”.

References

Alberti, M., Marzluff, J., Hunt, V.M., 2017. Urban driven phenotypic changes: empirical observations and theoretical implications for eco-evolutionary feedback. Philosophical Transactions of the Royal Society B: Biological Sciences 372(1712). https://​doi​.org/​10​.1098/​ rstb​.2016​.0029 Boyd, R., Richerson, P.J., 1988. Culture and the Evolutionary Process. University of Chicago Press. Brooks, J.S., Waring, T.M., Borgerhoff Mulder, M., Richerson, P.J., 2018. Applying cultural evolution to sustainability challenges: an introduction to the special issue. Sustainability Science 13, 1–8. https://​doi​.org/​10​.1007/​s11625​ -017​-0516​-3 Durham, W.H., 1991. Genes, Culture, and Human Diversity. Stanford University Press. Ehrlich, P.R., Raven, P.H., 1964. Butterflies and plants: a study in coevolution. Evolution 18(4), 586–608. Foxon, T.J., 2011. A coevolutionary framework for analysing a transition to a sustainable low carbon economy. Ecological Economics 70(12), 2258–67. https://​doi​.org/​10​.1016/​j​

Miquel A. Gual and Richard B. Norgaard

.ecolecon​.2011​.07​.014 Goddard, J., Kallis, G., Norgaard, R., 2019. Keeping multiple antennae up: coevolutionary foundations for methodological pluralism. Ecological Economics 165, 1–9. Gowdy, J.M., 1994. Coevolutionary Economics: The Economy, Society, and the Environment, Natural Resource Management and Policy. Kluwer Academic Publishers. Gowdy, J.M., 2021. Ultrasocial: The Evolution of Social Nature and the Quest for a Sustainable Future. Cambridge University Press. Gual, M.A., 2005. Políticas de promoción de la energía renovable: un modelo de análisis sistémico-coevolutivo [Renewable Energy Promotion Policies: A Coevolutionary Systemic Framework]. Universidad Pablo de Olavide. https://​rio​.upo​.es/​xmlui/​handle/​10433/​11839 Gual, M.A., Norgaard, R., 2010. Bridging ecological and social systems coevolution: a review and proposal. Ecological Economics 69, 707–17. Haider, L.J., Schlüter, M., Folke, C., Reyers, B., 2021. Rethinking resilience and development: A coevolutionary perspective. Ambio 50, 1304–12. Jumonville, N., 2002. The cultural politics of the sociobiology debate. Journal of the History of Biology 35, 569–93. Levin, S., 2005. Self-organization and the emergence of complexity in ecological systems. Bioscience 55(12), 1075–9. Lumsden, C.J., Wilson, E.O., 1981. Genes, Mind, and Culture: The Coevolutionary Process. Harvard University Press, Cambridge, MA. Miles, L.S., Carlen, E.J., Winchell, K.M., Johnson, M.T.J., 2021. Urban evolution comes into its own: emerging themes and future directions of a burgeoning field. Evolutionary Applications 14, 3–11. https://​doi​.org/​10​.1111/​eva​.13165 Moreno-Peñaranda, R., Kallis, G., 2010. A coevolutionary understanding of agroenvironmental change. A case-study of a rural community in Brazil. Ecological Economics 69, 770–78. https://​doi​.org/​10​.1016/​j​.ecolecon​.2009​.09​.010 Norgaard, R., 1981. Sociosystem and ecosystem coevolution in the amazon. Journal of Environmental Economics and Management 8, 238–54. https://​doi​.org/​10​.1016/​0095​ -0696(81)90039​-5 Norgaard, R., 1984. Coevolutionary agricultural development. Economic Development and Cultural Change 32(3), 525–46. https://​doi​.org/​ 10​.1086/​451404. Norgaard, R., 1989. The case for methodological pluralism. Ecological Economics 1, 37–57. Norgaard, R., 1994. Development Betrayed. The End of Progress and a Coevolutionary Revisioning of the Future. Routledge. Norgaard, R., 2021. Post economics: reconnecting reality and morality to escape the Econocene. Real-World Economics Review 22(96), 49–66. Ríos-Núñez, S.M., Coq-Huelva, D.,

Coevolution (socio-biophysical coevolution)  69 García-Trujillo, R., 2013. The Spanish livestock 2019. Evolution in the Anthropocene: informmodel: a coevolutionary analysis. Ecological ing governance and policy. Annual Review Economics 93, 342–50. of Ecology, Evolution, and Systematics 50, Saifi, B., Drake, L., 2008. A coevolutionary 527–46. https://​doi​.org/​10​.1146/​annurev​ model for promoting agricultural sustainability. -ecolsys​-110218​-024621 Ecological Economics 65, 24–34. https://​ doi​ Søgaard-Jørgensen, P., Folke, C., Henriksson, .org/​10​.1016/​j​.ecolecon​.2007​.11​.008 P.J.G., Malmros, K., Troell, M., Zorzet, A., Shiki, S., de Faria, S., Shiki, N., Lopes Rosado, 2020. Coevolutionary governance of antibiotic P., Aparecida Fernandes, E., 2013. Payment and pesticide resistance. Trends in Ecology and for ecosystem services markets and rural Evolution 35, 484–94. https://​doi​.org/​10​.1016/​j​ development in Brazil—a co-evolutionary .tree​.2020​.01​.011 approach. Journal of Environmental Science Spash, C.L., 2013. The shallow or the deep ecoand Engineering B2 (2013), 276–88. https://​doi​ logical economics movement? Ecological .org/​10​.17265/​2162​-5263/​2013​.05​.002 Economics 93, 351–62. https://​doi​.org/​10​.1016/​ j​.ecolecon​.2013​.05​.016 Søgaard-Jørgensen, P., Folke, C., Carroll, S.P.,

Miquel A. Gual and Richard B. Norgaard

12. Common property and environmental governance

erty rights are to be assigned, monitored, or sanctioned; Ostrom and Hess 2010). Common property rights grant control over resources to groups of users of the resources (local fisher, irrigator, forest groups) in the expectation that they will be able to coordinate with each other, build a regime, and avoid overexploitation or pollution problems within a certain territory. In many cases, these regimes are embodied in organizations, such as Water User Associations (WUAs), community-forest associations, or fishing cooperatives. In theory, common property regimes are an alternative to private and public property regimes (see section 12.2). In reality, however, all three “solutions” tend to coexist (see section 12.4). Common property theory explains the conditions under which coordination can emerge and endure over time (see section 12.3). Much of this theory speaks about local contexts and communities; however, scholars have also paid attention to large resource systems and the interactions between communities and other public and private actors (Fleischman et al. 2014; see section 12.5 below). Also, the theory has been criticized for its limited attention to the role of history and internal politics (Johnson 2004; Agrawal 1994), ignorance of the role of power in the effective implementation of property rights (Ribot and Peluso 2003), inattention to external threats to the regimes and social mobilization (Villamayor-Tomás and García-López 2018), and exclusive focus on the economic rationale behind the idea of natural resource use and coordination (García-López et al. 2017; see sections 12.6 and 12.7).

12.1 Introduction

This introduction to common property and environmental governance takes the standpoint of institutional ecological economics (Paavola and Adger 2005; see sections 12.1–12.5 below) and some of the criticisms addressed to it (sections 12.6 and 12.7). Institutional ecological economics studies environmental degradation as a problem of coordination among resource users, and governance as a way to promote said coordination. Coordination problems exist because natural resource users, and groups of them (including polluters), are interdependent in their use of the resources.1 For example, farmers and firms at local scales, or national governments at the international scale, are interdependent in the use they make of water from, for example, a shared aquifer; water use decisions that a farmer, firm, or national government makes will affect the amount or quality of water available to other farmers, firms, and governments, respectively (and vice versa). Interdependence becomes particularly evident when resources and services suffer from congestion. Complaints of urban drinking-water users to upstream farmers about water contamination or shortages become salient as water demand in cities or irrigation systems increases and/or water availability decreases (e.g., during dry seasons). In these kinds of contexts, environmental governance is understood as the rule, and norm-based solutions (e.g., devised by farmer associations, firms, local and/or national governments) promote coordination among actors. Property rights are rules and norms that define who can benefit from a resource and/or make decisions about how to use it. Property rights are at the basis of environmental governance because they specify who can use the resource, and who can say how to use the resource, among other aspects (Schlager and Ostrom 1992; Sikor et al. 2017).2 That said, property rights do not mean much if they are not embedded within an institutional regime (i.e., the set of rules that frame how the prop-

12.2

Common vs. private and public property regimes

Property right theory in the natural resource management context has traditionally distinguished three broad types of regimes. In private property regimes, each individual user of a resource system (e.g., a forest, a fisheries, or a river) has control over a parcel of that resource system (e.g., land tracts, fishing grounds, river tracts) or an amount of resource units that can be extracted from the system (a share of timber, fishes, water). Alternatively, common property regimes grant ownership of the entire resource system to the ensemble of users or user groups (i.e., no parceling), and public property regimes 70

Common property and environmental governance  71

grant it to a public authority. Each regime has, in turn, rules to ensure proper functioning of the regime. In private property right regimes, each owner decides how to use and manage their parcel of the resource, and rules usually specify how to solve conflicts or exchange rights among them. In public property regimes, the governments design rules to specify how the resource should be used, by whom, and when. In common property regimes, the ensemble of users or user groups entitled with the collective right need to self-organize to create those rules. Thus, contrary to the other regimes, common property regimes also require rules that articulate local collective decision-making among users.

12.3

Conditions for success in common property regimes

Private, public, and common property regimes can all fail (Acheson 2006). Common property regimes do not always emerge due to the inability of users to cooperate with each other. Also, common property regimes do not necessarily prevent environmental degradation if user groups are not concerned about it, if the communities lack important information about the conditions of the resource, or if their conservation efforts are jeopardized by external users or public authorities (Acheson 2006). At large, common property regimes are more likely to work if basic collective governance processes, such as collective choice, monitoring-sanctioning, and conflict-solving, perform satisfactorily (i.e., they ensure that rules exist, fit users’ interests and understandings of the resource, and are followed by the users). This is more likely to happen in the presence of certain biophysical conditions (e.g., small and predictable resource systems), social conditions (e.g., small, socially cohesive groups that have previous cooperative experience), and institutional conditions (e.g., existence of rules that ensure a proportional distribution of costs and benefits among resource users and respect ecological dynamics like seasonality in resource availability; Poteete et al. 2010; Cox et al. 2010).

12.4

Hybrid regimes

The theoretical distinction between common, public, and private property regimes applies only partially to real-life scenarios; public,

private, and common property rights coexist quite frequently. In many irrigation systems worldwide, for example, WUAs have the right to use and manage the water within their jurisdictions (common property regime), but not to sell the water to other associations or urban users, because this right stays with the government (Garces-Restrepo et al. 2007).3 Also, use or management rights in both private and common property regimes are usually conditioned to higher-level regulations about what are acceptable uses (i.e., which reflect the right of governments to manage certain aspects of resource). In the fishing context, for example, Individual Transferable Quota (ITQ) programs grant private use and management rights to individual fishers and firms, but these are usually conditional to compliance with endangered species regulations or international fishing agreements. Sometimes also, governments and local communities share management rights, which translates in the joint development of regulatory, monitoring, and/or conflict-solving mechanisms. In community-forest regimes in India, or irrigation systems in Nepal, for example, governments and communities jointly decide how monitoring and conflict-solving are to be carried out (Weiland and Dedeurwaerdere 2010; Frey et al. 2016). Similarly, in some WUAs and exceptional situations (e.g., during droughts) farmers are allowed to exchange their shares of water with each other as if they held them privately (Villamayor-Tomás 2014; Svensson et al. 2021); and it has been a relatively frequent practice in some fishing contexts that holders of ITQs “communalize” those quotas ad hoc to share the risk of bycatch (Holland and Jannot 2012; Zhou and Segerson 2016).

12.5

Common property in large-scale contexts

An increasingly relevant question in the advent of environmental problems like biodiversity loss and climate change is whether lessons from local and regional common property regimes apply to the global scale. At first regard, global environmental problems are quite different from local ones. The former are much less prone for exclusion, involve many more users, and are more complex both socially (more heterogeneous Sergio Villamayor-Tomás

72  Elgar encyclopedia of ecological economics

users) and biophysically (interlinked land, water, and atmospheric systems) than the latter. Also, change at global scales is particularly unpredictable, and “we have only one globe with which to experiment” (Stern et al. 2001; Ostrom et al. 1999, p. 282). On the other hand, research around regional and global resource systems, such as large marine protected areas, large forests, international rivers, or the ozone layer, has shown that, despite the distinct challenges, similar governance logics apply to both. Clearly defined boundaries and monitoring of resource conditions are crucial, and who wins and who loses in controlling resource use degradation is also important at both scales (Fleischman et al. 2014).

12.6

Property rights and the power to exercise them

An important critique to common property theory has been its relative ignorance of politics (i.e., the use of power to promote the interests of some user groups over others) and its impact on the endurance and effectiveness of common property regimes. From a politics lens, rights are just one among other sources of power that ultimately provide access to natural resources. “People may hold property rights to some resources without having the capacity to derive any material benefit from them” (Sikor and Lund 2009, p. 4) because they do not have the appropriate technology, financial, human, or knowledge capital, or connections with key authorities (Ribot and Peluso 2003). On the one hand, there are the politics internal to the regimes. As hinted in political ecology works, the presence of common property regimes does not ensure equity in the distribution of the resource among participants, nor the lack of conflict in case of inequalities. Differences in socioeconomic traits between social groups (e.g., differences in wealth or status across gender, ethnic, or class divides) can drive inequalities in the way the resource and decision-making power are allocated (Blaikie 2006; Agrawal and Gibson 1999); and privileged groups can use resource use and decision-making rules to protect their privileges and reproduce inequalities (Agrawal 1994). On the other hand, there are the politics external to the regimes. The effectiveness of common property regimes to promote Sergio Villamayor-Tomás

sustainable use of a resource within a territory may be threatened by the actions of external users (e.g., “roving bandits” in the fisheries sector or large timber companies in the forest context) or the imposition by governments of development and/or conservation policies that jeopardize the regime’s rules (Villamayor-Tomás and García Lopez 2021a). In response, user groups may mobilize and strike back via protests, alliances with user groups in other locations, and engage in party and high-level politics with governments to defend their regimes (Villamayor-Tomás and García-López 2018).

12.7

Common senses beyond scarcity

As hinted in the introduction, much of the justification for property rights from the institutional ecological economics perspective rests on the logic of scarcity and the need to control resource use in the advent of its degradation; however, evidence has also pointed to other, less economic, mechanisms through which common property regimes emerge and are sustained over time. The interest of Indigenous communities in controlling use may well have to do with certain understandings of their relationship with the environment and “mother Earth” (Escobar 2019; Barkin 2020) and religious beliefs (Hartberg et al. 2014). Similarly, the willingness of user groups to create common property regimes may be driven by certain ethics, for example, about the rights of nature (Dupuits et al. 2020); shared understandings that caring about people and nature go hand in hand (García-López et al. 2017; Clement et al. 2019); or postmaterial lifestyles (Gibson and Koontz 1998). Finally, a fair number of so-called commons movements and regimes have emerged worldwide in the last decades to counterbalance neoliberal and privatization policies both ideologically and practically (Villamayor-Tomás and García-López 2021b).

12.8

Concluding remarks

Common property theory emerged in the 1980s to complement prescriptions from previous theory about the suitability of private and public property governance to solve resource degradation problems. Common property theory soon became mainstream

Common property and environmental governance  73

and embodied in forest and irrigation decentralization policies and in community-based conservation programs worldwide. However, common property regimes (just like public or private regimes) are no panacea. From an institutional ecological economics perspective, their performance depends on the rules that organize basic governance processes like decision-making, enforcement, and conflict-solving. Thus, the question is not whether common property regimes are better or worse than public or private property regimes, but the conditions under which each can work best, alone or in combination with each other. Ideally, research on these conditions should also pay more attention to the integration of power dynamics and non-economic logics. Sergio Villamayor-Tomás

Notes 1.

See Ostrom et al. (1994) for a deeper discussion of how this interdependence is related with the non-excludable and depletable nature of shared natural resources (i.e., common pool resources). 2. As reviewed in Ostrom et al. (1994), natural resources that are depleted or easily congested and that are not easily excludable are called common pool resources (CPRs). This is different from common property regimes (which can also be seen in the literature as CPR). As noted in Ostrom and Hess (2010), CPRs do not need to be always governed via common property regimes. 3. As conceptualized by Schlager and Ostrom (1992), users that have the right to trespass a resource system may not have the right to harvest products from it; users that have the right to use the resource may not have the right to manage it or exclude others from using it; and having these rights does not necessarily involve having the right to sell the resource.

References

Acheson, James M. 2006. “Institutional Failure in Resource Management.” Annual Review of Anthropology 35: 117–34. Agrawal, Arun. 1994. “I Don’t Need It but You Can’t Have It: Politics on the Commons.” Pastoral Development Network 36: 36–55. Agrawal, Arun, and Clark C. Gibson. 1999. “Enchantment and Disenchantment: The Role of Community in Natural Resource Conservation.” World Development 27 (4): 629–49. https://​doi​ .org/​10​.1016/​S0305​-750X(98)00161​-2 Barkin, David. 2020. “Local Responses to Global Change: Community Alternatives for ‘Good Living’ in Latin America.” In V.I. Grover, A.B. Borsdorf, J. Breuste, P.C. Tiwari, and F.W.

Frangetto (eds), Impact of Global Changes on Mountains, 391–408. CRC Press. https://​ doi​ .org/​10​.1201/​b17963​-24 Blaikie, Piers. 2006. “Is Small Really Beautiful? Community-Based Natural Resource Management in Malawi and Botswana.” World Development 34 (11): 1942–57. Clement, Floriane, Wendy Harcourt, Deepa Joshi, and Chizu Sato. 2019. “Feminist Political Ecologies of the Commons and Commoning (Editorial to the Special Feature).” International Journal of the Commons 13 (1): 1. https://​doi​ .org/​10​.18352/​ijc​.972 Cox, Michael, Gwen Arnold, and Sergio Villamayor-Tomás. 2010. “A Review of Design Principles for Community-Based Natural Resource Management.” Ecology and Society 15 (4): 38. http://​www​.ecologyandsociety​.org/​ vol15/​iss4/​art38/​ Dupuits, Emilie, Michiel Baud, Rutgerd Boelens, Fabio de Castro, and Barbara Hogenboom. 2020. “Scaling Up but Losing Out? Water Commons’ Dilemmas between Transnational Movements and Grassroots Struggles in Latin America.” Ecological Economics 172 (June): 106625. https://​doi​.org/​10​.1016/​j​.ecolecon​ .2020​.106625 Escobar, Arturo. 2019. “Thinking-Feeling with the Earth: Territorial Struggles and the Ontological Dimension of the Epistemologies of the South.” In B. de Sousa Santos and M. Meneses (eds), Knowledges Born in the Struggle, 41–57. Routledge. https://​doi​.org/​10​ .4324/​9780429344596​-3 Fleischman, Forrest, Natalie Ban, Louisa Evans, Graham Epstein, Gustavo García-López, and Sergio Villamayor-Tomás. 2014. “Governing Large-Scale Social-Ecological Systems: Lessons from Five Cases.” International Journal of the Commons 8 (2): 428–56. https://​ thecommonsjournal​.org/​articles/​10​.18352/​ijc​ .416 Frey, Ulrich J., Sergio Villamayor-Tomás, and Insa Theesfeld. 2016. “A Continuum of Governance Regimes: A New Perspective on Co-Management in Irrigation Systems.” Environmental Science & Policy 66: 73–81. https://​doi​.org/​10​.1016/​j​.envsci​.2016​.08​.008 Garces-Restrepo, Carlos, Douglas Vermillion, and Giovanni Muñoz. 2007. “Irrigation Management Transfer: Worldwide Efforts and Results.” FAO Water Reports. Rome, Italy: Fod and Agriculture Organization of the United Nations. García-López, Gustavo A., Irina Velicu, and Giacomo D’Alisa. 2017. “Performing Counter-Hegemonic Common(s) Senses: Rearticulating Democracy, Community and Forests in Puerto Rico.” Capitalism Nature

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74  Elgar encyclopedia of ecological economics Socialism 28 (3): 88–107. https://​ doi​ .org/​ 10​ .1080/​10455752​.2017​.1321026 Gibson, Clark, and Tomas Koontz. 1998. “When ‘Community’ Is Not Enough: Institutions and Values in Community-Based Forest Management in Southern Indiana.” Human Ecology 26 (4): 621–47. https://​ doi​ .org/​ 10​ .1023/​a:​1018701525978 Hartberg, Yasha, Michael Cox, and Sergio Villamayor-Tomás. 2014. “Supernatural Monitoring and Sanctioning in Community-Based Resource Management.” Religion, Brain & Behavior 6 (2): 95–111. https://​doi​.org/​10​.1080/​2153599X​.2014​ .959547 Holland, Daniel S., and Jason E. Jannot. 2012. “Bycatch Risk Pools for the US West Coast Groundfish Fishery.” Ecological Economics 78 (June): 132–47. https://​doi​.org/​10​.1016/​j​ .ecolecon​.2012​.04​.010 Johnson, Craig. 2004. “Uncommon Ground: The ‘Poverty of History’ in Common Property Discourse.” Development and Change 35 (3): 407–34. Ostrom, Elinor, Joanna Burger, Christopher B. Field, Richard B. Norgaard, and David Policansky. 1999. “Revisiting the Commons: Local Lessons, Global Challenges.” Science 284 (5412): 278–82. https://​doi​.org/​10​.1126/​ science​.284​.5412​.278 Ostrom, Elinor, Roy Gardner, and James Walker. 1994. Rules, Games and Common Pool Resources. Ann Arbor: University of Michigan Press. Ostrom, Elinor, and Charlotte Hess. 2010. “Private and Common Property Rights.” Property Law and Economics 5: 53. Paavola, Jouni, and W. Neil Adger. 2005. “Institutional Ecological Economics.” Ecological Economics 53 (3): 353–68. https://​ doi​.org/​10​.1016/​j​.ecolecon​.2004​.09​.017 Poteete, Amy R., Elinor Ostrom, and Marco Janssen. 2010. Working Together: Collective Action, the Commons, and Multiple Methods in Practice. Princeton, NJ: Princeton University Press. Ribot, Jesse C., and Nancy Lee Peluso. 2003. “A Theory of Access.” Rural Sociology 68 (2): 153–81. Schlager, Edella, and Elinor Ostrom. 1992. “Property-Rights Regimes and Natural Resources: A Conceptual Analysis.” Land www​ .jstor​ Economics 68 (3): 249–62. http://​ .org/​stable/​3146375 Sikor, Thomas, Jun He, and Guillaume Lestrelin. 2017. “Property Rights Regimes and Natural Resources: A Conceptual Analysis Revisited.”

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World Development 93 (May): 337–49. https://​ doi​.org/​10​.1016/​J​.WORLDDEV​.2016​.12​.032 Sikor, Thomas, and Christian Lund. 2009. “Access and Property: A Question of Power and Authority.” Development and Change 40 (1): 1–22. https://​doi​.org/​10​.1111/​J​.1467​-7660​ .2009​.01503​.X Stern, Paul C., Thomas Dietz, Nives Dolsak, Elinor Ostrom, and Susan Stonich. 2001. “Knowledge and Questions After 15 Years of Research.” In Paul C. Stern, Thomas Dietz, Nives Dolsak, Elinor Ostrom, Susan Stonich, and Elke U. Weber (eds), The Drama of the Commons, 445–91. National Academy of Sciences. Svensson, Jesper, Yahua Wang, Dustin Garrick, and Xiaoping Dai. 2021. “How Does Hybrid Environmental Governance Work? Examining Water Rights Trading in China (2000–2019).” Journal of Environmental Management 288 (February): 112333. https://​doi​.org/​10​.1016/​j​ .jenvman​.2021​.112333 Villamayor-Tomás, Sergio. 2014. “Cooperation in Common Property Regimes under Extreme Drought Conditions: Empirical Evidence from the Use of Pooled Transferable Quotas in Spanish Irrigation Systems.” Ecological Economics 107 (0): 482–93. https://​doi​.org/​10​ .1016/​j​.ecolecon​.2014​.09​.005 Villamayor-Tomás, Sergio, and Gustavo García-López. 2018. “Social Movements as Key Actors in Governing the Commons: Evidence from Community-Based Resource Management Cases across the World.” Global Environmental doi​ Change 53 (November): 114–26. https://​ .org/​10​.1016/​J​.GLOENVCHA​.2018​.09​.005 Villamayor-Tomás, Sergio, and Gustavo A. García-López. 2021a. “Commons Movements: Old and New Trends in Rural and Urban Contexts.” Annual Review of Environment and doi​ .org/​ 10​ Resources 46 (1): 511–43. https://​ .1146/​annurev​-environ​-012220​-102307 Villamayor-Tomás, Sergio, and Gustavo García-López. 2021b. “Decommonization-Commonization Dynamics and Social Movements: Insights from a Meta-Analysis of Case Studies.” In P.K. Nayak (ed), Making Commons Dynamic: Understanding Change through Commonization and Decommonization, 255–82. Routledge. Weiland, Sabine, and Tom Dedeurwaerdere. 2010. “Change in Forest Governance in Developing Countries – in Search of Sustainable Governance Arrangements.” International Journal of the Commons 4 (2): 683. https://​doi​.org/​10​.18352/​ IJC​.258 Zhou, Rong, and Kathleen Segerson. 2016. “Individual vs. Collective Approaches to Fisheries Management.” Marine Resource Economics. MRE Foundation. https://​doi​.org/​ 10​.1086/​685384

13. Complex social-ecological systems

(Holmes, 2001), and ecosocial systems (Waltner-Toews et al., 2003). Theoretically, the SES concept has been raised to deal with this separation (Berkes et al., 2003; Westley et al., 2002), to stress mutual dependency among the subsystems, and to emphasize the same weight for all of them (Berkes, 2017; Berkes et al., 1998). Contributions of Ecological Economics to this conceptualization have been widely recognized (Folke et al., 2016). The debate has moved from isolation/partial connection vs. clear links to a discussion about the nature of the links, in particular, the discussion between those considering a hierarchy of systems and those presenting the two subsystems co-evolving at the same hierarchical level. The former presenting a larger ecological system in which the social system is embedded, with the scale of human activities on the biosphere as the ultimate sustainability problem (Daly, 1990; Goodland and Daly, 1996), giving space for different metabolic approaches to the sustainability problems (Giampietro et al., 2014, 2009; Haberl et al., 2011). The latter present mutual dependencies among systems that must be managed to reach sustainability (Kallis and Norgaard, 2010; Norgaard, 1981). SESs are open systems. Taking into account the nature of the interaction with the environment, systems can be characterized as closed (i.e. limited interchanges with the environment), isolated (i.e. no interchanges with the external environment), or open (i.e. with interchanges with the external environment; von Bertalanffy, 1968, 1950). Since social systems and ecosystems depend on external flows to stabilize their own state, SESs have been characterized as far from thermodynamic equilibrium (Prigogine, 1955; Schneider and Kay, 1994), and consequently, radically open systems. Expanding this approach, the idea of SESs as complex systems appears (Holland, 1995; Levin, 1999). As complex systems, some characteristics can be stressed. First, is non-linearity (i.e. inputs in complex systems could have unexpected outputs; Glansdorff and Prigogine, 1971). As a consequence, the system behaviour is nearly uncertain, therefore, it cannot be directly predicted using a traditional cause and effect analysis (Kay et al., 1999; Kovacic and Di Felice, 2019). Also, it presents self-organization (i.e. coherent patterns of relationships internally structured

According to the most quoted definition, read between the lines in Berkes et al. (1998, 4), Social-Ecological Systems (SESs) are complex, integrated systems in which humans are part of nature (humans-in-nature). Under diverse names, this concept has been widely used in many disciplines, including Ecological Economics (Farrell and Luzzati, 2013; Ostrom, 2005, 2009). As a consequence, numerous definitions of SES can be found, each one stressing different aspects of interest, but sharing the idea of humans and nature as a whole (Colding and Barthel, 2019). First of all, SESs are systems. The concept of system appears throughout social and natural sciences and emerged out of a recognition of the limitations of traditional mechanistic and reductionist traditional approaches of the Western science (Hammond, 2017). Systems have been defined as groups of material or abstract elements interacting between them and an environment as a whole (von Bertalanffy, 1968). It is possible to conceptualize ecological systems or ecosystems, constituted by the species interacting between them and their environment at different scales (Odum, 1953; Tansley, 1935), and social systems as processes of interaction between actors (Parsons, 1951) or as a network of communication between people (Luhmann, 1995, 1982), but also systems of humans-in-nature, or SESs. One of the common factors in the various definitions of SESs is the integration of these subsystems in a larger system (SESs as integrated systems). Nowadays it seems obvious, but the conceptual and methodological integration of humans and nature has not been the norm in science, and different forms of separation have been hegemonic in most of the scientific fields, at least from the Enlightenment (Glacken, 1967; White, 1967). But at the same time, different approaches to deal with the interplay between these systems have been developed, often emphasizing a specific dimension, for example, human–environment systems (Scholz and Binder, 2011), coupled human–nature systems (Alberti et al., 2011; Liu et al., 2007), socio-ecological systems 75

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and developed over time; Kay and Boyle, 2008; Maturana and Varela, 1980; Nicolis and Prigogine, 1989). Thus, complex systems are hierarchically structured (Allen and Giampietro, 2014; Allen and Starr, 1982), therefore, smaller subsystems are nested in larger subsystems. And finally, unexpected properties can emerge among hierarchical levels (Holland, 1998; Kauffman, 1993). This acknowledgement of the open and complex nature of SESs and its thermodynamic consequences in Ecological Economics has been one of the main arguments presented in the debate with supporters of the possibility of perpetual growth, and their modern decoupling and circular views of the economic process (Georgescu-Roegen, 1971; Giampietro, 2019; Kapp, 1976). Moreover, SESs have frequently been presented as a particular category of complex systems: the complex adaptive systems (CAS) (Levin et al., 2013; Preiser et al., 2018). Organizational patterns emerge in response to changes in the constitutive system and its environment so that SESs have the capacity to adapt to such changes (i.e. adaptive capacity). In the context of the sustainability debates, this complex adaptive nature of SESs has been often cited. Thus, it has been suggested as the cause for the failure of optimization or efficiency static approaches when applied to single variables or to maximize the output over short time frames in simple theoretical models to prescribe universal solutions (Benessia et al., 2016; Holling and Meffe, 1996; Mayumi and Giampietro, 2006; Saltelli and Giampietro, 2017). Consequently, in conceptual terms, the academic attention has moved from optimization and efficiency in the human exploitation of nature to other perspectives based on different SES abilities. Especially, resilience, or the capacity to absorb, adapt, or transform with change (Gunderson, 2000; Holling, 1973), creates a stability domain in which the system maintains its identity. But adaptability (i.e. the capacity of human actions to manage resilience in order to maintain a certain stability domain) and transformability (i.e. the capacity to shift current development pathways into other ones, or to create novel pathways, allowing different stability domains) are also important (Folke et al., 2010; Walker et al., 2004). At the same time, certain heuristic tools, like the adaptive cycle (Gunderson and Pedro L. Lomas

Holling, 2002; Holling, 1986), have been proposed to explain these changes. To articulate the different concepts associated to the sustainability of SESs, the need for a framework has been highlighted (Ostrom, 2005). A framework would be useful in providing a common set of relevant variables to be used in the design of data collection instruments, fieldwork, and the analysis of findings about the sustainability of complex SESs (Ostrom, 2009). Different frameworks have been proposed to operationalize the SES concept (Binder et al., 2013), including: (1) Driver-Pressure-State-Impact-Response (DPSIR) provides a framework for categorizing a sustainability problem based on the causal chain of drive r-pressure-state-impact-response indicators (European Environment Agency, 1999), and describes SESs in terms of social system actions on the environment and feedbacks of policies. (2) Ecosystem Services (ES), in which ecosystems are conceived as providers of services for human well-being. This framework was originally conceived to highlight the natural basis of human well-being, although different forms of nature commodification have prevailed (Gómez-Baggethun et al., 2010), leading to the concept of Nature’s Contribution to People (NCP), which is related to the original ES concept, expanding the framework to the non-instrumental values and claiming plurality of values (Pascual et al., 2017). (3) In Earth System Analysis (ESA) human activities are presented as drivers of change (Schellnhuber et al., 2005). The linkages between these subsystems are represented as flows of matter and energy quantified by using different coupled models (Flato, 2011). (4) A metabolic framework (MF) is based on the metabolism concept applied to the relationships between ecological and social systems. At least two main proposals can be identified (Gerber and Scheidel, 2018): on one hand, the Material and Energy Flow Accounting (MEFA) or social metabolism, based on the quantification of stocks and flows of matter and energy circulating from

Complex social-ecological systems  77

nature to society at different scales (Haberl et al., 2011); on the other hand, Multi-Scale Integrated Assessment of Social and Ecosystem Metabolism (MuSIASEM), based on the assessment of interdependent social and ecosystem metabolisms (Giampietro et al., 2014, 2009). (5) The Social-Ecological Systems Framework (SESF) is a multilevel, nested framework for analysing outcomes at the SES level (Ostrom, 2009, 2005; Partelow, 2018). SESF tries to understand the SESs’ functioning, and the development, implementation, and transformation of SESs towards sustainability goals. It is based on the division of an SES into four core subsystems (resource system, resource unit, governance systems, and users) that affect each other as well as linked social, economic, and political settings and related ecosystems. Each core subsystem is made up of multiple second-level variables, chosen according to the particular questions and temporal and spatial scale under study. (6) Robustness is an engineering concept very similar to resilience for intentionally designed systems. With the main goal of preserving the robustness of the SESs, main entities and their relationships are characterized: resources, resource users, public infrastructure providers, public infrastructures, institutional rules, and external environment (Anderies et al., 2019, 2004). Despite the original intention of dealing with the separation of natural and social systems, it has been argued that these frameworks are often unbalanced and loaded to the natural side. Thus, many authors claim that relevant social dimensions, such as power, competing value systems, human agency, or normative issues, are largely ignored by merely considering the ecological understanding of the SESs’ characteristics and dynamics (Armitage et al., 2012; Cote and Nightingale, 2011; Davidson, 2010; Stone-Jovicich, 2015). Also, there is a general agreement on the fact that the complex and changing nature of SESs requires new ways of management to reach societal sustainability goals. To this purpose, different forms of adaptive manage-

ment (Holling, 1978; Walters, 1986) have been proposed. In general, adaptive management is about increasing knowledge (not only scientific knowledge, but also other relevant types of knowledge) and, at the same time, allowing the management to proceed despite the inherent uncertainty associated to SESs. New self-organized formal and informal structures of governance emerge in these adaptive management processes (i.e. adaptive governance structures; Folke et al., 2005). Some authors have proposed the adaptive management to reinforce SESs’ resilience, as an ultimate goal, in a context of incomplete knowledge and urgent action required despite inherent uncertainty (Allen and Garmestani, 2015; Holling, 1978; Walters, 1986). However, it has been recognized that this type of adaptive management, called active adaptive management, is only desirable and possible in a context of high uncertainty and controllability (Allen and Gunderson, 2011); others propose a different type of adaptive management sustained in the view of Ecological Economics as a post-normal science (Funtowicz and Ravetz, 1994), based on the role of scientists as narrative providers (quantitative story telling) instead of predictors (scientific evidence; Saltelli and Giampietro, 2017), and their participation as equals with an extended peer community in the process of updating the social metabolism while remaining operational in a context of environmental change and uncertainty. Pedro L. Lomas

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14. Consumption

housing, or transport studies. It is characterised by its own toolbox of approaches and the degree of its institutionalisation is visible in journals, collaborative scholarly groups (European Roundtable on Sustainable Consumption and Production [ERSCP], Sustainable Consumption Research and Action Initiative [SCORAI]), encyclopedias (Reisch & Thøgersen, 2015), or research agendas (Mont, 2019), as well as the standard inclusion of consumption-related entries in broader handbooks on ecological economics or sustainable development (Røpke, 2015; Spash & Dobernig, 2017). But what is the role of ecological economics in consumption research, then? Asking this question, we do not intend to etch out a specific area that is supposed to be the sole domain of ecological economists or the domain of solely ecological economics. Yet, it should be possible to identify areas and questions to which ecological economics can add in particular, given its specific theoretical perspectives and methodological approaches. We believe that the above question probably can be answered best via a list of research foci. In our view, core themes in (sustainable) consumption research to which ecological economics can contribute in particular are:

14.1 Introduction

The main focus in research on consumption in ecological economics is the (un)sustainability of consumption. This focus has resulted from the recognition that overconsumption of resources in the Global North and among the economic elites in the Global South is a major contributor to the depletion and destruction of the Earth’s ecosystems, while others still consume too little (Martin et al., 2021; Wiedmann et al., 2020). Since the 1970s and especially the early 1990s, scholars have explored questions of how to improve the sustainability of consumption. The resulting complexity and diversity of consumption research makes it almost impossible to provide an overview of the field. In the following, we therefore try to sketch the field via a discussion of three (interacting) aspects: a) the evolution of its structures, b) empirical developments, and c) core research themes.

14.2

The evolution of the field’s structures

● systems of provisioning and consuming; ● the governance of consumption and the power of actors and ideas; and ● consumption, limits, and quality of life.

Research on consumption and environment predates the establishment of ecological economics as a field. From the early 1960s, environmentalists questioned the environmental impacts of consumption, and in the 1970s the first research contributions emerged, particularly from the marketing field of consumer behaviour and from energy studies in the wake of the energy crises. But the real take-off came in the late 1980s and early 1990s, alongside the Earth Summit in 1992, when the topic also became part of ecological economics. Over the following decades, then, consumption and environment became an almost intractable area of research. In the mid-2000s, when the Internet Encyclopedia of Ecological Economics was written, much research on sustainable (and unsustainable) consumption could be seen as related to ecological economics, and it was possible to provide a reasonable overview (Røpke, 2005/2006). Today, however, sustainable consumption has become its own multidisciplinary research field, with extensions into other fields such as energy,

Before presenting a brief overview of research developments on these research foci, however, let us sketch relevant empirical developments that have accompanied and interacted with the evolution of the field.

14.3

Empirical developments

With the Rio summit in 1992, consumption officially became part of the global political agenda. Agenda 21 explicitly calls for the adoption of sustainable consumption patterns (United Nations [UN], 1993). In the following years, international organisations, such as the UN Environment Programme, the Organisation for Economic Co-operation and Development, and the European Union, started to pursue a range of initiatives in this area. They investigated trends in and impacts of consumption patterns and tried to facilitate the development of new or extended 81

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policy strategies and regulatory frameworks (Fuchs & Lorek, 2005). By the time of the Johannesburg summit in 2002, however, a lack of political action became visible, and ideas for a new framework emerged (UN/ World Summit on Sustainable Development [WSSD], 2002). When this was eventually adopted at the Rio+20 Conference in 2012, it was soon overshadowed by the Sustainable Development Goals (SDGs), adopted by the UN in 2015. Indeed, SDG 12, “Ensure sustainable consumption and production patterns”, simultaneously marks the acknowledgement that sustainable consumption is an important global development goal and relegates it into a chorus of 17 goals. Overall then, we have been witnessing around 30 years of activities addressing sustainable consumption in the political arena. Knowledge has increased, regulations have improved the ecoefficiency of products, consumers have been provided with more information and sometimes enticed with economic incentives to make ecologically superior choices. But unfortunately, we are not on track towards sustainable consumption. In the same period, the ecological footprint of consumption has systematically gotten worse (Wiedmann et al., 2020). In the Global North, consumers use up even more resources than before, and in the Global South, a new economic elite has emerged following the same patterns (Myers & Kent, 2004). Resource savings induced by technological innovation have been overpowered by rebound effects and general consumption growth. For instance, households may use more efficient appliances, but they are also using more of them more of the time with increasingly faster replacement rates (European Environment Agency, 2015). Home insulation has improved, but residential space per person keeps increasing; and more efficient cars have become available, but mobility increases, also in terms of more plane travel (Eurostat n.d.). To some observers, the COVID-19 pandemic appeared to offer a chance for households to re-evaluate their consumption patterns. However, after months of lockdown, many individuals appear more eager than ever to get back on the consumption train, and the experience of effective government intervention does not seem to be transferred to the sustainability issue. Doris Fuchs and Inge Røpke

14.4

Core research foci

Why have we achieved so little in halting the destruction of the Earth’s ecosystems caused by overconsumption? The three core research foci in sustainable consumption research laid out below seek answers to this question and investigate strategies for change. Again, we discuss them separately for ease of understanding, even though they speak to each other, of course. 14.4.1 Systems of provisioning and consuming With the biophysical perspective of ecological economics, it is obvious to explore the environmental impacts of consumption. Some studies highlight the impacts of particular products and services, applying life cycle analysis, but most studies focus on broader categories of consumption and apply less-fine-grained approaches, such as energy accounting, materials flow accounting, ecological footprint, human appropriation of net primary production, carbon footprint, and water footprint. These approaches often apply environmentally extended input–output analyses (increasingly taking trade into account) and demonstrate the high resource intensity of food, mobility, and housing, as well as different impacts of various sub-categories (Tukker et al., 2010, 2018). Combining these data with data on the consumption patterns of different income groups, the environmentally detrimental impacts of high-income groups become visible (Oswald et al., 2020). In addition, studies apply data on energy and material flows to expose how resources, through processes of unequal exchange in biophysical terms, are transferred from poor to rich countries and enable the high consumption of the rich (Dorninger et al., 2021). The biophysical explorations beg the question of the drivers and mechanisms behind the ever-growing consumption. As a starting point, there are two basic preconditions that have made high levels of consumption possible in the Global North (Røpke, 2015). First, the availability of cheap and high-quality fossil energy has been decisive for the increasing labour productivity that results in the abundance of material goods (Ayres et al., 2003; Haberl et al., 2011). Second, the large inequalities both within and between countries are key to consumption growth. For instance, the extremely low wages in mining,

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agriculture, and industrial sweatshops in poor countries provide rich countries with cheap products, while local elites may profit from the trade arrangements (Schor, 2005). The mechanisms of transfer and the underlying power relations have emerged over a long span of time and continually change (Hornborg et al., 2007). These basic preconditions coevolve with the capitalist growth engine, where competition drives business to innovate processes and products, and thus increases labour productivity and tempts consumers with novel offers, supported by advertising, deferred payment, planned obsolescence, and so on. While the capitalist growth engine certainly drives consumption, it still leaves the key question of why consumers are willing to play their part (Røpke, 1999). Some scholars emphasise the role of consumption in the symbolic communication between people in relation to status seeking or identity formation (Brekke et al., 2003). Others argue that, particularly in relation to environmental impacts, the ordinary, inconspicuous consumption related to everyday use of water, heating, electricity, and so on is more important (Shove & Warde, 2002). Much consumption is not really a matter of choice, for instance, when people experience technological or institutional lock-in effects (Southerton et al. 2004). Moreover, consumption is integrated in everyday practices in ways that imply that consumption is not even considered as such (Røpke, 2009; Godin et al., 2020; Shove et al., 2012; Welch & Warde, 2015). Since social practices are seen as shared and influenced by social and material conditions, the approach involves a critique of the traditional governance focus on influencing individual consumer choices. Instead, it calls for politicians to take on responsibility for shaping the conditions for social practices (Spurling et al., 2013). Similar policy recommendations emerge from studies on socio-technical systems for provisioning and consuming, whether they emerge from the tradition of Fine and Leopold (Fine et al., 2018; Mattioli et al., 2020), or from the tradition of transition theory in combination with practice theory (McMeekin & Southerton, 2012; Hargreaves et al., 2013). In general, these approaches fit very well with the systems orientation of ecological economics.

14.4.2 Governance and power Governance-focused inquiries into the (un) sustainability of consumption have investigated the nature of historical and ongoing governance efforts and the determinants of governmental choices in approaches, targets, and instruments. Here, scholars have demonstrated that almost all (inter-)governmental activities have focused on the efficiency of consumption, and on encouraging and enabling consumers to buy more efficient products or services (Fuchs & Lorek, 2005). With the exception of the “Consumption Opportunities” report (UNEP, 2001), hardly any relevant publication by (inter-)governmental actors in the first decades inquired into possibilities of “overconsumption” and “misconsumption”. Even today, few targets of SDG12 show much ambition, and those with some ambition remain limited by a focus on relative rather than absolute progress (Bengtson et al., 2018). This neglect of “strong sustainable consumption governance” (i.e. governance addressing levels and fundamental patterns of consumption, as well as the predominant choice of informational or incentive-based policy instruments) can be explained with the power of (especially economic) actors benefitting from the status quo as well as related ideational and material structures (Fuchs et al., 2016, 2021b; Martiskainen et al., 2019). Actors such as the International Chamber of Commerce, the World Business Council for Sustainable Development, or the World Federation of Advertisers lobbied for such a narrow approach to sustainable consumption and a focus on educating and encouraging consumers via reports and positions papers at the international level, and their counterparts pursued the same strategies at the national level (Fuchs & Lorek, 2005). More fundamentally, a triad of myths about the salvatory potential of technological innovation, the ability of markets to contribute to societal problem solving, and the existence and power of consumer sovereignty served to hide the need for a different type of approach (Fuchs et al., 2021a). These myths resonated with dominant narratives about the value of deregulation and privatisation and found their reflection in the ecological modernisation perspective and its emphasis on the potential for win–win solutions, voluntary self-regulation, and consumer information Doris Fuchs and Inge Røpke

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via certification or labelling schemes. We still recognise these myths and their underlying normative foundations in today’s focus on nudging and prosuming in research and policy. Even though scholars and activists highlighted the defects of “private governance” (Clapp, 1998; Gibson, 1999) and of the responsibilisation of consumers (Maniates, 2001; Princen, 1997) early on, the myths’ fit with dominant politico-economic paradigms, not least capitalism, allowed them to prevail. Indeed, the consumerist society has become accepted as a natural phenomenon, hiding how political support and incessant advertising fostered its development to begin with (Isenhour, 2017; Kasser, 2016). Thus, ideas related to the necessity of limits to growth and consumption are creeping onto the political agenda only very belatedly, influenced by protest movements such as Fridays for Future and the increasingly dire warnings of the Intergovernmental Panel on Climate Change and Scientists for Future. 14.4.3 Consumption, limits, and quality of life So perhaps the insight that overconsumption is causing the breaching of planetary and societal boundaries is slowly becoming speakable in political arenas. Related to the individual level, this overconsumption can be defined as consuming so many resources that the chances of other individuals living now or in the future to have access to a quantitative and qualitative minimum of resources necessary to be able to live a good life are being hurt. Defined in this way, overconsumption implies both consumption minima and maxima and relates them to the normative goals of a good life for all, planetary health, and justice. It is most prominently captured in the concept of consumption corridors (Fuchs et al., 2021a), but also informs related ideas of 1.5°C Lifestyles (Institute for Global Environmental Strategies [IGES], 2019) or Doughnut Economics (Raworth, 2017). Such a perspective on consumption, then, also reveals that the notion of sufficiency needs to be thought of in the sense of two forms of enough: enough for the individual and enough for all (Spengler, 2016). Moreover, it allows us to differentiate between needs (limited and satiable), wants (potentially unlimited and non-satiable), and satisfiers (means to satisfy needs and wants), and the differences in their Doris Fuchs and Inge Røpke

legitimacy and societal responsibilities for their fulfilment (Di Giulio & Defila, 2020). Minimum consumption standards are a well-known notion in welfare states. Maximum consumption standards appear more controversial, even though they also exist all around us already (Fuchs et al., 2021a) and are compatible with liberal notions of freedom (Gumbert & Bohn, 2021). Most fundamentally, consumption limits run counter to capitalist dynamics of accumulation (Pirgmaier, 2020) and ideas of a correlation between increasing consumption and improvements in the quality of life. However, once needs are met, contentedness with one’s quality of life depends on a multitude of individual and societal aspects (Jackson, 2017). Scholars and activists highlight the physical, emotional, and cognitive benefits of downshifting, slow or simple living, reductions in working hours, or time wealth (Geiger et al., 2021). Specifically, our communities are also decisive for our welfare, and they, in turn, are shaped by environmental quality, social (in) equalities, the health and education systems, infrastructure, and so on. To the extent that economic growth and increasing levels of consumption foster a deterioration of societal conditions, their positive contribution to quality of life has to be in doubt. Of course, we have to be careful not to take the “double dividend” (Spargaaren, 1997) of reducing consumption in terms of improved quality of life and planetary well-being for granted or become moralistic in the pursuit of maximum consumption standards. It is easier to imagine a society with less consumption and a higher quality of life (less stress, debt, or cluttering, less “defensive” costs, improved collective conditions) than to achieve it. Individually, at least, people make other choices in practice. Yet as long as a decoupling of economic growth and resource consumption is not in sight (Parrique et al., 2019), the size of the economy has to be limited in the interest of future generations (as well as other species). To allow for others living now or in the future to live a good life, the affluent have to reduce their consumptive footprint. The “efficiency revolution” needs to be accompanied by a “sufficiency revolution” among the richest one-fifth of the world population, and as societies, we need to talk about conditions of a good life, about what is enough, and about maximum consumption standards (Fuchs et al., 2021a; Gough, 2020).

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14.5 Outlook

Where do we stand after a good three decades of consumption-focused research in ecological economics (and beyond)? It would seem that we already know a lot of what we need to know about the ecological (and social) impacts of overconsumption. Yet, the translation of this knowledge into political action is not taking place. Are the barriers to sustainable consumption impossible to penetrate? Clearly, consumption is embedded in social and cultural life and thus efforts to reduce consumption levels necessarily interfere with a lot of what we have gotten used to. Also, such efforts necessarily imply challenging powerful actors benefitting from the status quo and dominant ideas that shaped development for many decades if not centuries. How can we bring about the necessary transformation of consumption levels and patterns, despite these obstacles? When will politicians and citizens be willing to do what is necessary to protect planetary boundaries and social well-being? These questions throw light on an area of research that continues to require particular attention: the role of capitalism and especially financial forces and logics. They also underline the relation to environmental activism and questions of science communication or more fundamentally the role of science in society. Thus, ecological economics, in cooperation with other fields, will need to continue to contribute to pursuing knowledge and action in this regard. Doris Fuchs and Inge Røpke

References

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in I. Røpke Consumption and Environment – Ecological Economic Perspectives (PhD thesis), University of Roskilde, Denmark, 12–27. Røpke, I. 2009. Theories of practice – new inspiration for ecological economic studies on consumption. Ecological Economics 68(10):2490–97. Røpke, I. 2015. Sustainable consumption: transitions, systems and practices. In J. Martinez-Alier and R. Muradian (eds.), Handbook of Ecological Economics. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing, 332–59. Schor, J. 2005. Prices and quantities. Ecological Economics 55(3):309–20. Shove, E., Pantzar, M., Watson, M. 2012. The Dynamics of Social Practice. London: Sage. Shove, E., Warde, A. 2002. Inconspicuous consumption. In R. Dunlap, F. Buttel, P. Dickens, and A. Bijswijt (eds.), Sociological Theory and the Environment. Lanham, MD: Rowman & Littlefield, 230–50. Southerton, D., Chappells, H., Van Vliet, B. (eds.). 2004. Sustainable Consumption. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Spargaaren, G. 1997. The Ecological Modernization of Production and Consumption. Wageningen: Landbouw Universiteit. Spash, C., Dobernig. K. 2017. Theories of (un)sustainable consumption. In C. Spash (ed.), Routledge Handbook of Ecological Economics. London: Routledge, 203–13. Spengler, L. 2016. Two types of “enough.” Environmental Politics 25(5):921–40. Spurling, N., McMeekin, A., Shove, E., Southerton, D., Welch, D. 2013. Interventions in Practice. Sustainable Practices Research Group, Lancaster University. Tukker, A., Cohen, M., Hubacek, K., Mont, O. 2010. The impacts of household consumption and options for change. Journal of Industrial Ecology 14:13–30. Tukker, A., Giljum, S., Wood, R. 2018. Recent progress in assessment of resource efficiency and environmental impacts embodied in trade. Journal of Industrial Ecology 22(3):489–501. United Nations (UN). 1993. Earth Summit: Agenda 21. New York: UN. United Nations Environment Programme. 2001. Consumption Opportunities. Paris: UNEP. UN/World Summit on Sustainable Development. 2002. Plan of Implementation. New York: UN. Welch, D., Warde, A. 2015. Theories of practice and sustainable consumption. In L. Reisch and J. Thøgersen (eds.), Handbook of Research on Sustainable Consumption. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing, 84–100.

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15. Cost shifting, competition and economic structure

structure of capitalism creates the conditions of competition that incentivise business enterprises to gain at others’ expense. Social costs are then incurred by those who are not responsible for, and typically not benefiting from, the economic activity creating these negative consequences. In Kapp’s work, social costs are understood as more than simply financial, and can involve a range of harm, including physical, psychological, and social impacts for humans and damages to non-human life and ecosystems. Social-ecological economics incorporates Kapp’s cost shifting theory within the structures of biophysical reality to explain the modern ecological crisis. Economies are understood as instituted processes of social provisioning that can take many forms. Most importantly, social reproduction of economies is connected to material and energy flows (i.e. a social metabolism). The basic physical laws of conservation mean that the use of energy and materials creates waste that inevitably goes back into either land, air, or water. Impacts result from both the scale of economic activity and the qualities of products and waste created during production and consumption processes. The pervasiveness of social costs is then explicable in terms of modern economic structures of capital accumulation (growth) and price-making markets (Spash 2020). Understanding the structure of an economic system is necessary for determining the causal mechanisms creating social costs. A prime environmental example is the increase of pollution associated with industrialisation, which has meant widespread and continuous impositions of costs on others by both private enterprise and the public sector. Cost shifting facilitates the avoidance of pollution control costs. For example, a popularised claim is captured by the phrase ‘the solution to pollution is dilution’. When put into practice, this has meant engineering the dispersion of pollutants in ever larger bodies of air, water, and soil. Rather than making pollutants harmless, this created a range of unexpected consequences, including pollutants appearing in food chains, being bio-accumulated, transported over long distances, and pushed onto future generations (i.e. ignorantly shifting costs spatially and temporally). The expanding scale, number, and harmful qualities of new pollutants (e.g. DDT, toxic chemicals, pesticides, insecti-

15.1 Introduction

Modern industrialised economies have created extensive environmental degradation which economic theory explains as social costs. Mainstream microeconomists regard social costs as anomalies arising in capitalist market economies due to pricing failures that result in private profit/utility maximising decisions diverging from social welfare maximisation. They are typically regarded as easily correctible via price adjustments. Ecological economics has highlighted an additional macroeconomic cause of social costs due to the drive for economic growth (Daly 1992), with an implicit focus on capitalist economies that has recently become more explicit (Spash 2020, 2021b). More generally, a combination of incentivising competition for personal (or in-group) gain and accumulation of capital are implicated in why modern economies, from corporate capitalist to state socialist, are associated with such a large range of ever-increasing environmental problems. Yet, despite the evidence, economists have persisted in regarding social costs in general, and pollution in particular, as a minor annoyance rather than a major systemic problem. Thus, until the attention given to climate change, environmental harm could be, and was, almost totally ignored by most economists, with only a minority of sub-disciplinary specialists showing any concern. Where social costs like pollution appeared, they were conceptualised as something outside the normal running of economic activity—an anomaly, a diseconomy, a disservice (Spash 2021a). In contrast to seeing social costs as anomalies, the institutional economic theory of cost shifting, developed by Karl William Kapp (1963/1978), explains how economic actors are incentivised to deliberately and intentionally create environmental impacts and harm. Conventions, norms, and formally sanctioned rules and regulations provide the institutional structure enabling the operation of any economy. Specifically, the institutional 88

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cides, plastics, radioactive waste, acidic deposition, and greenhouse gases) is then correlated with development equated with economic growth, and associated with new technologies, including exploitation of new forms of energy and materials, and hi-tech militarisation to secure supply chains. Kapp’s (1950) work on social costs predates the general rise of environmental concern and provides insights into the economics of pollution some decades before environmental economics was even established as the neoclassical economist’s response. His work poses a substantive and foundational challenge to this orthodox thinking, and an original theoretical approach. While there has been some recent attention (e.g. Berger 2017; Elsner et al. 2007; Gerber and Steppacher 2012), most economists have failed to recognise the distinctiveness of his cost shifting explanation or its importance (Leipert 1986; Swaney and Evars 1989). Indeed, rather than adopting his institutional and social-ecological economic critique, many ecological economists have instead tended to unthinkingly adopt mainstream externality theory. As will be shown, this employs a caricatured and incorrect interpretation of Arthur C. Pigou’s ideas (Aguilera Klink 1994; Spash 2021a) that lacks his insights, let alone those of Kapp. We therefore start by placing the concept of social costs in historical context as an idea arising in economic thought, and explain its development in the writings of Marshall, Pigou, and Coase. We then turn to the alternative cost shifting theory of Kapp and briefly outline some of the implications this has for environmental policy. While cost shifting is, as will be explained, broadly definable across both social and ecological phenomena, in this chapter we mainly have in mind the damages and harms associated with environmental impacts.

15.2

A brief history of social costs

Neoclassical economics describes markets as self-regulating structures that tend towards an equilibrium of production (supply) and consumption (demand) that is meant to be optimal in terms of the efficient use of resources. This is predicated upon assuming that an economic actor is rational if they seek

to maximise their own self-interest. A competitive market system is then meant to allow such actors to operate as agents free from interference by others so that they can obtain the best, hedonistic outcomes for themselves. Kapp (1963/1978) reviewed various past critiques and qualifications of these claims, including those from: classical economists (Smith, Malthus, Ricardo), the historical school (Compte, Schmoller), socialist economists (Marx, Sisimondi, Hobson, Lange), welfare economists (predominantly Pigou), and institutional economists (Veblen, Clark). His review reveals how economic systems based on private interests have long been recognised as leading to outcomes that diverge from those of broader society. In contrast, the modern theory of social costs as externalities arose in the late 1950s and early 1960s with a heritage from Alfred Marshall, and ignores this more extensive critical literature (Spash 2021a). In terms of neoclassical economics, Marshall’s investigation of external economies and diseconomies was an early acknowledgement that the decision-making processes of capitalist private enterprise economies fail to include all the relevant costs and benefits to society of its activities. He especially raised the topic of ‘external economies’ as the unpaid benefits accruing to firms from investments made by other private firms or the public sector. They might result from generally expanding markets, access to a trained workforce, higher standards of health, education, or culture (Kapp 1970: 841). Marshall’s concept of (dis)economies captures issues related to how the social and material conditions evolve that enable any given social provisioning system to operate. This evolutionary aspect of his theory was, he thought, complimentary to a mechanistic analysis (e.g. market supply and demand). However, the economics profession became dominated by mechanistic static equilibrium analysis to the exclusion of all else, and explicitly (e.g. Stigler, see Kapp 1970) rejected Marshall’s more evolutionary ideas. Economists have generally referenced Pigou (1920) with founding their notion of external costs—not a term he actually employed (Spash 2021a). In Pigou’s work there is some correspondence with Marshall in as far as Pigou’s concepts of social services and disservices define divergences between Clive L. Spash and Amelia Fuselier

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marginal social net product and marginal trade net product. Pigou also maintains the claim that welfare-increasing government corrections are feasible to maximise the national dividend (i.e. gross domestic product), taken as the means for welfare increases. However, as Kapp (1970: 338) argues, Marshall’s theory concerns the dynamic adjustment of economic structure to increase production/ wealth, not the management and regulation of specific welfare-harming activities, which appear in the work of Pigou. Economists later attributed to Pigou the assertion that all social costs could be calculated in monetary terms for every single unit individually (i.e. marginal social cost estimation). Neoclassical economists, and their environmental economic specialists, developed this into a theory where an optimal level of pollution can be determined as the most efficient using the methods of social cost– benefit analysis. In something of a contradiction to economists’ faith in ‘free markets’ having the ability to produce efficient prices, government policy is necessary to ‘get the prices right’ by employing taxes based on calculating shadow prices. Pigou himself explicitly concluded that this was impossible (see Spash 2021a). Yet this failed to prevent his concept of social disservices being appropriated and his policy recommendations converted into the misnomer of a ‘Pigouvian’ tax—a simplistic optimal price adjustment calculated to exactly match marginal social costs (i.e. monetised harm and damages). Pigou’s own thesis not only denies the general use of ‘the measuring rod of money’, but (contra the erroneous claims of Coase 1960) also addresses itself to a wide range of corrective measures including legal contract, direct regulation, and the rule of law (Aguilera Klink 1994; Spash 2021a). Pigou’s economic proposals are generally reformist, but by current standards appear quite radical. In his later work he promoted socialist policies and public ownership (Pigou 1937/1949). Distinguishing and advocating the advantages of a socialist economy over one based on private enterprise was popular at the time, even amongst neoclassical economists (e.g. Lange and Taylor 1938). That is, a basic structural reform of capitalist economies was deemed necessary. Indeed, this highlights how the common causal mechanisms of social costs were recognised to be embedded in the structure Clive L. Spash and Amelia Fuselier

of competitive growth economies. Stigler typifies those attempting to counter such a realisation in defence of capitalism and price-making markets and, in his case, claims that there is no common causal mechanism underlying the divergence of private from social costs (see Spash 2021a: 10). The pervasiveness of social costs evidenced by Pigou’s examples was then employed to argue that their diversity meant their causes were ad hoc. In fact, the primary concern of Chicago School economists, like Stigler, was an ideological defence of private enterprise capitalism, against claims of its structural inadequacies, and mainstream economic price theory against its foundational flaws due to the pervasiveness of unmonetised (and non-monetisable) social costs. As neoliberalism spread, economists inspired by neo-Austrian economic ideology (e.g. the Chicago School) attributed social costs to a lack of private property rights. They saw no need for the neoclassical economists’ interventionist pricing policy, which implied centrally planned prices set by government. Instead, they promoted the idea that whether polluters or victims bore the costs of pollution did not matter, as long as the costs became part of market-based transactions and were ‘freely’ negotiated as such. Rather than correcting prices, all that was needed was well defined and legally enforceable private property rights. This is typically referred to as the ‘Coase Theorem’, although that is a bastardisation of Coase’s actual work created by Stigler and referred to by Coase himself as ‘the infamous Coase Theorem’ (see Spash 2021a: 98). In modern environmental policy, cost shifting has been made into a virtue by instituting private property rights via offset and permit markets that seek the least-cost means of allowing pollution and environmental destruction to continue. In the 1960s, a few economists recognised the all-pervasive nature of pollution and briefly made explicit the revolutionary implications for economic theory (e.g. Kneese et al. 1970). However, they quickly fell back into line with the neoclassical orthodoxy that became environmental economics, and no revolutionary paradigm shift occurred. What became generally termed ‘externalities’ remained, for economists, a correctible market failure in an otherwise perfectly efficient economic system. Mainstream economists have persisted in claiming their ability

Cost shifting, competition and economic structure  91

to prescribe optimal levels of harm and damages, and faith in market price adjustment (e.g. via taxes or tradable permits) as the solution. Whether believing that economic agents should pay government taxes on every atom in existence or private property rights could be established for every atom in existence, economists appear united in their understanding that pollution should be conceptualised as something ‘external’ to normal (capitalist) economic activity. Only in the work of Kapp was this dominant position substantively rejected, its scientific inadequacies exposed, and an alternative theoretical explanation offered.

15.3

Cost shifting and the cumulative causation of social harm

In developing his theory of social costs, Kapp employs aspects of evolutionary thinking and institutional analysis. There is some correspondence with Marshall’s (dis)economies, but Kapp goes well beyond this rather limited conceptualisation of the issue. In fact, Kapp (1963/1978) conceptualises social costs differently from both Marshall and Pigou. Social costs are the consequences of productive processes that could be avoided but are instead deliberately created because an economic actor benefits by doing so. They are by definition harms, damages, expenses, and generally negative impacts shifted on to ‘others’: private or public, individual or communal, human or non-human. Examples of social costs include pollution of air, water, and soil; work-related morbidity and mortality; impacts of natural resource extraction, species loss, and ecosystem degradation; and density of urbanisation (Kapp 1963/1978). Swaney and Evers (1989) build further on the cost shifting approach by denoting consequences far removed from the original production process and a chain of impacts on others (second to nth party effects), such as the psychological stress of performance pressure. Competition incentivises gain over others, which can be achieved by producers making others bear the costs of their business enterprises. Profit maximisation dictated by competitive market economies implies a systemic tendency to minimise costs. As Kapp

(1963/1978: 14) states: ‘the basic causes of social costs are to be found in the fact that the pursuit of private gain places a premium on the minimisation of the private costs of current production’. Profit maximising actors are incentivised to reduce and avoid the costs they need to take into account as private decision-makers (Kapp 1963/1978). Such actors regard their own interest as above the common good or social well-being. Core to their economic activities is the ability to shift costs (Kapp 1963/1978: 14). Outcompeting others for personal gain is the price-making market game. Prices are then the outcome of a power struggle involving the ability to shift costs and achieve mark-ups. Market power is evident in actualised cost shifting and the dynamic of shaping social institutions to progressively facilitate further cost shifting (Kapp 1950; Swaney and Evars 1989). Such a process is evident in the regulatory capture of government agencies and the increasing role of government in service of corporate (as opposed to public) interest. Success in the competitive marketplace is then defined by the ability to control the process of cost shifting to make it work in one’s own favour. This creates a cascade of costs shifting to labour and other producers. Kapp (1963/1978) identifies the structure of competitive capital accumulating economies as creating social exploitation and environmental degradation that destabilise economic systems (i.e. causing social and ecological crises). What then arises is the tendency of capitalism to cause societal and institutional change that undermines its own stability and continued operation. While noting the relevance of Schumpeter and Marx in this context, Kapp himself drew on the institutional economics of Veblen and incorporated his concept of cumulative causation. This highlights the principle ‘that social processes are marked by the interaction of several variables both “economic” and “noneconomic” which in their combined effects move the system away from a position of balance or equilibrium’ (Kapp 1963/1978: 25). Cost shifting practices impact on interdependent biophysical, social, and economic systems, leading to a chain of reinforcing consequences (Kapp 1970). Feedback processes reinforce causal mechanisms already activated and cause further rounds of causal reinforcement. Under cirClive L. Spash and Amelia Fuselier

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cular and cumulative causation, rather than an initial change calling forth countervailing forces, it brings forth supporting changes that move the system further in the same direction. This explains how actual dynamic processes of economic growth reproduce environmentally destructive processes that are neither efficient nor socially beneficial (Kapp 1963/1978: 272). Initially Kapp’s focus was on private enterprise in the form of (American) capitalism. However, he came to realise that the social organisation of production as a process of competitive cost minimisation could be institutionalised under state-planned or entrepreneurial/corporate capitalist economies. Thus, the organisation of business units in the USSR appeared just as problematic. Indeed, any economic system could institutionalise the causal mechanism of cost shifting in production that led to the pervasiveness of social costs. In light of this, he renamed his major book on the topic to be more inclusive, replacing ‘private’ with ‘business’ enterprise (Kapp 1963/1978). However, on the basis of Kapp’s insight that competitive social structures incentivise cost shifting behaviour, the phenomenon is far from limited to business enterprises or production. Wherever seeking competitive advantage for gain over others is institutionalised, cost shifting can be expected. For example, regional governments competing with each other for tourism, corporate business, jobs, urban expansion, or other revenue-raising and growth opportunities may engage in lowering or avoiding regulatory standards for social and environmental protection (i.e. creating social costs). In terms of public policy responses, Kapp (1969) explains that social costs cannot be captured by the traditional accounting practices of social cost–benefit analysis. Some costs may be reflected in the monetary expenditure of others, for example, domestic pollution abatement equipment, such as household water filtration systems. However, many social costs are intangible, unquantifiable, and simply cannot be converted into a meaningful money metric. Kapp (1963/1978) gives numerous examples, including physical and psychological health, morbidity and mortality, and irreversible resource depletion. More recent environmental concerns (e.g. ecosystem health, biodiversity loss, climate change) have simply added to the list of things that are impossible to Clive L. Spash and Amelia Fuselier

reduce to a monetary value. Yet economists, such as Stern and Nordhaus on climate change and Dasgupta on biodiversity, have persisted in claiming their ability to apply social cost– benefit analysis to determine optimal levels of harm to others, and/or correct market failures with the ‘true’ prices (Keen 2021; Spash 2002, 2007; Spash and Hache 2021). The numerous problems with monetary calculation of social costs have become ever more evident as ecological crises have expanded (e.g. complex causal relationships, strong uncertainty over impacts, effects distant in time and space, plural and incommensurable values). The idea that marginal damage functions could be estimated is pure fiction, and even point estimates prove highly restricted and contestable. Kapp (1965) stresses the necessity for explicit judgement over (plural) values to imbue the assessment of gains and losses with meaning. He advocates the establishment of social minima required to achieve human objective needs (Kapp 1963/1978, 1965, 1972). This entails thorough scientific assessment combined with democratic and discursive methods (Kapp 1974). He died before substantively addressing how to incorporate plural and conflicting values in public policy and decision processes. Cumulative causation also poses the difficulty of determining the core causal mechanisms when confronting complex causal chains and open system dynamics. For social-ecological economists, these are urgent interdisciplinary research challenges.

15.4 Conclusion

The exclusion of the unmonetised impacts of productive activities from decision-making based on financial indicators is a commonly understood aspect of social costs. A straightforward conclusion is that unregulated competitive economic systems fail to maximise social welfare. Neoclassical economists then take the divergence of private from social costs to justify government intervention to adjust prices according to their social impacts, which are implicitly assumed to be treatable as monetary equivalents. This monetisation, let alone the optimal internalisation of externalities, was recognised by Pigou as impossible, and so taken to justify a wide range of regulatory approaches.

Cost shifting, competition and economic structure  93

Despite this, the predominant response to social costs today is embedded in neoclassical economic ideas of capitalist markets and prices as efficient resource allocators, even using Pigou’s name to support the idea of optimal taxes. On top of this, neo-Austrian ideological commitment to the private entrepreneur as core economic actor has encouraged spreading private property rights to establish new trading opportunities as ‘the solution’. Private gain at public expense has expanded as a result. The required public policies for addressing the ecological crisis (e.g. all-pervasive pollution, climate change, mass extinction of species) have nothing to do with creating new markets or getting the prices right! In contrast, Kapp’s concept of cost shifting offers a realist, institutional explanation of social costs that is consistent with the theoretical insights of social-ecological economics. This recognises value pluralism and incommensurability as making social costs (i.e. various forms of harm and damages) irreducible to a monistic monetary metric. Prices are the outcome of power processes and not freely determined objective efficient resource costs. Indeed, the prevalence of pollution highlights how all prices would need to be adjusted by government intervention to follow the logic of internalising externalities, or basically that all prices are ad hoc subject to which social costs are regulated and how or whether they are taken into account. Kapp’s cost shifting theory highlights the need for alternative institutional arrangements to price-making markets, competition, business enterprise, and self-interest. In addition, Kapp explains how dynamic processes of cumulative causation within competitive economic structures incentivise cost shifting practices, which then increase over time. The damages from pollution, for example, are not accidents or anomalies but rather inherent within such economic structures. Market capitalism is one form of economic system where cost shifting is prevalent, but not the only one, and Soviet business enterprise proved just as problematic, as has Chinese state capitalism. Cost shifting can only be addressed by changing the economic structure of social provisioning away from those exploitative forms that have become dominant today.

Ecological economists must then seek to identify the causal mechanisms of cost shifting and implement policies to counter them. For Kapp this involved determining and enforcing social minima and developing democratic participatory processes to do so. More fundamentally, his understanding identifies the need for both a paradigm shift in economic thought and a systemic change in actual economies. Clive L. Spash and Amelia Fuselier

References

Aguilera Klink, F. (1994). Pigou and Coase reconsidered. Land Economics, 70 (3): 386–90. Berger, S. (2017). The Social Costs of Neoliberalism: Essays on the Economics of K. William Kapp. Nottingham: Spokesman. Coase, R.H. (1960). The problem of social cost. Journal of Law & Economics, 3 (October): 1–44. Daly, H.E. (1992). The steady-state economy: alternative to growthmania. In: Daly, H.E. (ed), Steady-State Economics: Second Edition with New Essays. London: Earthscan, 180–94. Elsner, W., Frigato, P., and Ramazzotti, P. (eds). (2007). Social Costs and Public Action in Modern Capitalism: Essays Inspired by Karl William Kapp’s Theory of Social Costs. London: Routledge. Gerber, J.-F., and Steppacher, R. (eds). (2012). Towards an Integrated Paradigm in Heterodox Economics. Basingstoke: Palgrave Macmillan. Kapp, K.W. (1950). The Social Costs of Private Enterprise. New York: Shocken. Kapp, K.W. (1965). Economic development in a new perspective: existential minima and substantive raitonality. Kyklos, 18 (1): 49–79. Kapp, K.W. (1969). On the nature and significance of social costs. Kyklos, 22 (2): 334–47. Kapp, K.W. (1970). Environmental disruption and social costs: a challenge to economics. Kyklos, 23 (4): 833–48. Kapp, K.W. (1972). Social costs, neo-classical economics, environmental planning: reply. Social Science Information, 11 (1): 17–45. Kapp, K.W. (1974). Environmental Policies and Development Planning in Contemporary China and Others Essays. The Hague: Mouton & Co. Kapp, K.W. (1978 [1963]). The Social Costs of Business Enterprise. Nottingham: Spokesman. Keen, S. (2021). The appallingly bad neoclassical economics of climate change. Globalizations, 18 (7): 1149–77. Kneese, A.V., Ayres, R.U., and d’Arge, R.C. (1970). Economics and the Environment:

Clive L. Spash and Amelia Fuselier

94  Elgar encyclopedia of ecological economics A Materials Balance Approach. Washington, D.C.: Resources for the Future. Lange, O., and Taylor, F.M. (1938). On the Economic Theory of Socialism. Minneapolis: University of Minnesota Press. Leipert, C. (1986). Social costs of economic growth. Journal of Economic Issues, 20 (1): 109–31. Pigou, A.C. (1920). The Economics of Welfare. London: Macmillan. Pigou, A.C. (1949 [1937]). Socialism Versus Capitalism. London: Macmillan and Co. Spash, C.L. (2002). Greenhouse Economics: Value and Ethics. London: Routledge. Spash, C.L. (2007). The economics of climate change impacts à la Stern: novel and nuanced or rhetorically restricted? Ecological Economics, 63 (4): 706–13. Spash, C.L. (2020). A tale of three paradigms: realising the revolutionary potential of eco-

Clive L. Spash and Amelia Fuselier

logical economics. Ecological Economics, 169 (March): 1–14. Spash, C.L. (2021a). The contested conceptualisation of pollution in economics: market failure or cost shifting success? Cahiers d’Économie Politique/Political Economy Papers, 79 (1): 85–122. Spash, C.L. (2021b). ‘The economy’ as if people mattered: revisiting critiques of economic growth in a time of crisis. Globalizations, 18 (7): 1087–1104. Spash, C.L. and Hache, F. (2021). The Dasgupta Review deconstructed: an exposé of biodiversity economics. Globalizations, May: 1–24. Swaney, J.A. and Evars, M.A. (1989). The social cost concepts of K. William Kapp and Karl Polanyi. Journal of Economic Issues, 23 (1): 7–33.

16. Critical materials If we look at the definition of critical, we might encounter “of the greatest importance” (Cambridge Dictionary) or “extremely important” (Britannica). Both have in common that they focus on the importance of something in a given context, but they can also be related to a situation that has the potential to become disastrous. These definitions can be extrapolated to the present regarding some elements that are needed in advanced technologies, such as renewable energies, or in electric and electronic devices. Whether a certain material is critical or not depends, among many other things, on the moment in which it is being evaluated. For instance, during the Iron Age, to have a constant and stable supply of iron minerals was crucial for the development and maintenance of societies. As for the definition of critical materials, even if some regions have built their own definitions based on factors that we will define shortly, there is a certain consensus that the term refers to a series of materials whose availability, or rather the lack of it, could put the economy at risk. To put it briefly, they are materials that the region needs for different goals (mainly technological) but for which a constant supply cannot be assured. The criticality and availability of the raw materials that society needs is not something new. Although it may seem like something from the 21st century, we can find numerous examples in the recent past. Any material can become critical if the demand exceeds supply, and this situation can be illustrated through tungsten supply and demand, an element that was essential to manufacture ammunition and weapons during World War II. In other occasions, the limitations are related to the source of the element itself. Such is the case of gallium, which is a by-product generated during the processing of bauxite, and whose availability depends on the extraction and refining of the latter. The United States has a long tradition related with analyzing the materials that are critical concerning security interests, which started in the early 20th century. Different lists of critical materials have been developed over time, according to the different historical moments, to assure supply for national 95

processing and manufacturing companies to meet defense needs. The Defense National Stockpile Center was created after World War II to address future shortfalls in material availability, and it monitors more than 160 minerals. Following the Federal Strategy to Ensure a Safe and Reliable Supply of Critical Minerals in 2022, a list of 50 critical materials for the US was published. This is not an exception; many regions have also published equivalent lists, taking into account factors such as concentration of extraction, dependence on imports, use in different sectors, economic importance, environmental issues, political stability of the main producing countries, and so on. In the case of the European Union (EU), the concern about the supply of raw materials began in 1975. In 2008, the Raw Materials Initiative was created, and one of its tasks is to prepare periodic reports to identify raw materials that are critical for the EU. The first report was published in 2010 and, since then, there have been several updates. Below we show the list published in 2020, which includes a total of 30 materials that are considered critical, some of them being cobalt, lithium, gallium, or rare earth elements. As stated before, the supply of raw materials is becoming a global issue, and Canada and Australia, among other countries, have established similar lists (Table 16.1). Some of the elements in this table are widely used in the renewable energy sector. For instance, niobium, critical according to all these studies, is present in large quantities in certain types of wind turbines. Besides, all electric and electronic devices present in mobile phones, tablets, and even electric vehicles demand rare earth elements, cobalt, gallium, germanium, and tantalum, among others. As mentioned, one of the many items that makes a material critical is the concentration of extraction—in other words, reliance on imports. Some of the elements labeled as critical, such as gallium, niobium, rare earth elements, or cobalt, are mainly extracted from one territory. For instance, in 2020, at the world level, almost 97 percent of gallium and 58 percent of rare earth elements were produced in China. Additionally, 91 percent of niobium came from Brazil and 68 percent of cobalt from the Democratic Republic of the Congo.

96  Elgar encyclopedia of ecological economics Table 16.1 Materials and elements included in the critical raw material lists for selected countries   

Europea

Australiab

Canadac

United Statesd

Japane

Aluminum

x (in bauxite)

  

x

x

  

Antimony

x

x

x

x

x

Arsenic

  

  

  

x

  

Barite/barium

x

  

  

x

x

Beryllium

x

x

  

x

x

Bismuth

x

x

x

x

x

Borate

x

  

  

  

x

Cesium

  

  

x

x

x

Chromium

  

x

x

x

x

Cobalt

x

x

x

x

x

Coking coal

x

  

  

  

  

Copper

  

  

x

 

 

Fluorspar/fluor

x

  

x

x

x

Gallium

x

x

x

x

x

Germanium

x

x

x

x

x

Graphite

x

x

x

x

  

Hafnium

x

x

  

x

x

Helium

  

x

x

  

  

Indium

x

x

x

x

x

Lithium

x

x

x

x

x

Magnesium/magnesite

x

x

x

x

x

Manganese

  

x

x

x

x

Molybdenum

  

  

x

  

x

Natural rubber

x

  

  

  

  

Nickel

  

  

x

x

x

Niobium

x

x

x

x

x

Platinum group elements

x

x

x

x

x

Phosphate rock/phosphorus

x

  

  

  

  

Potash

  

  

x

  

  

Rare earth elements

x

x

x

x

x

Rhenium

  

x

  

  

x

Rubidium

  

  

  

x

x

Scandium

x

x

x

x

  

Selenium

  

  

  

  

x

Silicon metal/silicon

x

  

  

  

x

Strontium

x

  

  

  

x

Tantalum

x

x

x

x

x

Tellurium

  

  

x

x

x

Thallium

  

  

  

  

x

Tin

  

  

x

x

  

Titanium

x

x

x

x

x

Tungsten

x

x

x

x

x

Uranium

  

  

x

  

  

Vanadium

x

x

x

x

x

Zinc

  

  

x

x

  

Zirconium

  

x

  

x

x

Notes: a European Commission (2020). Communication from the Commission to the European Parliament, The Council, the European Economic and Social Committee and the Committee of the Regions. Critical Raw Materials Resilience: Charting a Path Towards Greater Security and Sustainability. COM/2020/474 final.

Alicia Valero, Guiomar Calvo, and Antonio Valero

Critical materials  97 b Australia Government (2019). Australia’s Critical Mineral Strategy. Department of Industry, Innovation and Science. Australian Trade and Investment Commission. c Natural Resources Canada (2021). Canada’s Critical Minerals Lists. Available at: http://ww​w.nrcan.gc​.ca/critic​ alminerals. d US Geological Survey (USGS) (2022). 2022 Final List of Critical Minerals. USGS, Department of the Interior. e Japan Organization for Metals and Energy Security (JOGMEC) (2019). Relevant Arguments for Formulating a New International Resource Strategy. Available at: https://www​.meti​.go​.jp/shi​ngikai/ene​ cho/shigen​_nenryo/sekiyu​_gas/pdf/010​_03​_00​.pdf.

At first glance, in a globalized market, this should not represent an issue. Nonetheless, the imposition of export quotas in some countries is increasingly frequent, causing the global supply of some elements to be compromised. To this we must add that the extraction of certain raw materials is concentrated in only one or two regions of the world. The processing facilities where the elements are extracted from the minerals are also located in a few of them. Such is the case for rare earth elements and China, where disruptions of supply could also compromise global availability. Another crucial issue regarding materials availability is the average concentration of elements in the mines, the ore grade, which influences the energy needed for the extraction process and the amount of waste rock generated. As the ore grade decreases in a mine, more energy is needed to obtain the same amount of ore as before, and it could reach a point where the process is no longer beneficial from an economic or environmental point of view.

Last, volatility of material prices can also influence critical materials and affect the whole supply chain. Mining exploration and exploitation is an activity for which large-scale investment projects are usually required. This process can last up to 10–20 years, depending on the commodity; this implies that it is a sector that is not able to react quickly enough to short-term changes in price and demand. As history has demonstrated several times, any material can become critical according to its demand and availability, or to the technology needed to supply said demand with the necessary speed. Historically, our world has depended on iron and copper. Now we rely on rare earth elements to manufacture permanent magnets, lithium and cobalt for electric vehicles, and tellurium for solar panels. Who knows what materials will be needed in the future, but, like so many times in the past and present, depending on their availability and demand, they could certainly be considered critical. Alicia Valero, Guiomar Calvo, and Antonio Valero

Alicia Valero, Guiomar Calvo, and Antonio Valero

17. Degrowth 17.1

The degrowth hypothesis

17.2

Degrowth’s origins

world is either a madman or an economist” (quoted in Olson 1973). In the 1960s and ’70s, some academics began to warn that Earth’s limits might impose the end of growth. Nicholas Georgescu-Roegen, in his 1971 book The Entropy Law and the Economic Process, explained the difference between stocks like oil and flows like solar power, and argued that as the stocks of fossil fuels deplete, economies would have to downsize to run on the flows of renewable energy. In the 1972 Limits to Growth report, Donella Meadows and colleagues described modeling scenarios in which economic growth produces its own demise via resource exhaustion and pollution. Predictions that the supply of petroleum would soon peak and then decline, seemingly supported by the oil crises of the 1970s, convinced some experts that growth would be over shortly. Around the same time, renegade economists began to theorize how modern economies might be made to function without growing. They advocated for bringing the zero-growth future into being. Herman Daly, in his 1973 book Toward a Steady-State Economy, devised a policy package for maintaining stable, sustainable levels of resource use and GDP. Published that same year, E.F. Schumacher’s Small is Beautiful promoted human-scale technologies and economies geared toward “enoughness.” Ivan Illich—a priest, not an economist—came out with a similar book, Tools for Conviviality, also in ’73; in it, he argued for technologies that empower ordinary people to act autonomously, and thus communities to control their lives without expert authorities. Radical greens in Europe at the time wanted to end grow-or-collapse capitalism and replace it with something like ecosocialism. These voices, in their calls for collective self-limitation, were precursors to the degrowth movement. The term “degrowth” did not yet exist. Yet André Gorz had begun to use décroissance, its French equivalent, in this ecological-economic sense. In 1979, professors Jacques Grinevald and Ivo Rens titled their French translation of collected Georgescu-Roegen essays Demain le Décroissance (“Tomorrow, Degrowth”) with the author’s consent. Radical environmentalists in France adopted the word as their slogan in the ensuing decades. Serge Latouche, an academic critic of Western-style

Degrowth means equitably downscaling the economy. It entails shrinking wealthy societies’ throughputs of materials and energy in ways that prioritize justice and well-being. Rich people and rich countries consume more resources than Earth can sustainably provide and generate more wastes than it can safely assimilate. Ecological economists argue that their resource use should be curtailed while ensuring everybody’s basic needs are met. Degrowth thus requires political change. Slowing down societies’ metabolisms will likely decrease gross domestic product (GDP), the total monetary value of everything an economy produces in a year. Economies will have to be restructured such that they can decelerate justly, without the suffering that recessions bring. But the degrowth hypothesis is not that downsizing rich economies will, on its own, facilitate a transformation toward egalitarianism and sustainability. It is, conversely, that improving social and environmental conditions may well necessitate, or even generate, a reduction in production and consumption. If everybody were guaranteed shelter without having to pay rent to landlords, for example, people would probably choose to work less, and might get by purchasing fewer things with the extra time off. A climate-saving moratorium on new fossil fuel extraction would, likewise, almost certainly constrain aggregate economic output in the short run. Degrowth is about reorganizing the economy to meet people’s needs regardless of what happens with GDP. Scholars of economics have long speculated about worlds beyond growth. John Stuart Mill (1848) imagined a “stationary state” in which culture flourishes once everyone’s material desires are satisfied. John Maynard Keynes (1932) thought that by the early 21st century the “economic problem” would be solved, and people would work just 15 hours per week. Kenneth Boulding once said something like, “Anyone who believes that exponential growth can go on forever in a finite 98

Degrowth  99

development, elaborated décroissance into an alternative to capitalism (see Latouche 2009). The concept emerged in English in 2008, when the scholar-activist collective Research & Degrowth organized the first of an ongoing series of biennial international conferences on degrowth. The degrowth literature has since flourished. Ecological economists show that green growth is not happening and not likely. Historians chronicle how the concept of growth emerged and came to dominate politics. Anthropologists conduct research in communities living without growth. Feminist authors demonstrate that caring for and about each other underlies diverse notions of the good life. Those who research and organize movements for environmental justice not only lay bare how rich countries’ growth begets conflicts in the places where materials are extracted and wastes discharged, they also tell stories of communities defending their traditional lifeways against the growth machine’s onslaught. Activists and academics have co-created the set of ideas that comprise degrowth. Those who prophesied that external constraints would curb growth were wrong, at least as yet. But it is not that the world became more resource efficient. The extraction and use of materials grew hand in hand with GDP. The few categories of materials whose global use has declined are nearly all substances that have been banned, such as asbestos and mercury. The resources that were supposed to run out simply did not. Technological change kept resources from running out. Green Revolution seeds and chemical fertilizers skyrocketed global food production and the ecological devastation it inflicts. Fracking and tar sands extraction have expanded the reserves of economically recoverable oil far beyond what can be burned without triggering catastrophic climate breakdown. This would probably not surprise Georgescu-Roegen, Meadows, or Boulding. Fifty years ago, they were already more worried about the economy’s tailpipe than its fuel tank. Today’s problem is damage, not scarcity. Scholars of degrowth call for “societal boundaries” to avoid transgressing planetary thresholds beyond which lie an altered and inhospitable Earth system (Brand et al. 2021). Instituting such limits is more a question of politics than of policy or science, at this point.

These limitations need not limit well-being. Degrowth’s proponents argue that democratically agreed-upon limits can open space for alternative and diverse conceptions of the good life and how to pursue it. Growth, for them, stands in the way of improved welfare. The literature on degrowth brings together critiques of growth from diverse disciplines and distant perspectives.

17.3

Critiques of growth

Organizing society around the pursuit of GDP growth is absurd. If one were to pay half of a city’s unemployed people to go out and break all the windows in town in the morning and the other half to replace them in the afternoon, that would increase GDP in the short run. If a hurricane breaks all the windows, same result: fixing them stimulates the economy. GDP counts costs, not benefits. It fails to capture the aspects of good living that are not bought and sold. And compound exponential growth—growing the economy by a consistent percentage—quickly shoots to infinity. Gross world product, or global GDP, is 30 times bigger than it was in 1900; at 3 percent growth, it will multiply tenfold again by the end of the 21st century. Even if the notion of physical limits to growth in the short run is debatable, the ecological critique of growth holds. Gross world product and total material extraction correlate as tightly as any variables outside of a laboratory. Humans already take a quarter of the productivity of all land on Earth (Krausmann et al. 2013). A global economy that is ten times bigger may, implausibly, require ten times as many tons of materials to be mined and harvested. The expansion of extraction causes social conflicts as well, because people live in and rely on the environments where resources are taken and pollution is dumped. Growth impedes mitigating climate change, too, since GDP and greenhouse gas emissions tend to rise and fall together. This is because fossil fuels still provide 84 percent of the world’s energy (BP 2020). Some countries, by switching to low-carbon energy, have managed to grow their economies while reducing emissions, but not at the pace needed to avert runaway warming (Haberl et al. 2020; Parrique et al. 2019). In climate models, the only scenarios that limit warming to 1.5 degrees Celsius without Sam Bliss and Giorgos Kallis

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relying on non-existent and risky “negative emissions technologies” are those in which global energy demand decreases dramatically (Grubler et al. 2018; Keyßer and Lenzen 2021). Transforming a smaller energy system to run on low-carbon fuels is easier than transforming a larger and growing one. Degrowth also follows in the tradition of critiquing growth’s ability to deliver well-being. There is little evidence that increasing income increases the happiness of nations (Easterlin et al. 2010). Having more money than others does correlate with greater subjective well-being, but this status-based phenomenon does not mean growth can make a population better off: one person ascending in rank means someone else is descending. Once everyone has a Mercedes-Benz, owning one can no longer confer status. Nor is growth necessary to ensure human health and longevity. Life expectancies have increased with improvements in hygiene and medicine, especially as fewer people have died in infancy and childhood. In poor countries that allocate modest resources to public health, people live just as long as in rich countries, in some cases longer. Excess economic production causes perhaps as many deaths today as lack. Outdoor air pollution kills over 4 million people each year (WHO 2021). Overly rich diets, physical inactivity, and overuse of tobacco and alcohol are major risk factors for the leading causes of death worldwide: heart disease and stroke (WHO 2020). Many ecological economists reckon that growth’s costs now exceed its benefits in wealthy countries. Daly (2014) calls this uneconomic growth because growth is then a net bad. If the things people need to be happy and satisfied—food, water, shelter, love, work, rest, entertainment, and so on—are not entirely substitutable between them, then at some point more production and consumption will make people worse off because it comes at the expense of time to care for each other and a healthy environment from which to live. We can’t eat entertainment. Research on degrowth goes beyond these ecological and economic critiques of growth. Growth is not just GDP increasing; it is the material manifestation of capital accumulating. It accelerates life, overworks and overstimulates us. It alienates us from each other and the rest of the mesh of existence. Sam Bliss and Giorgos Kallis

Growth has become the objective of modern societies. It distracts attention from tackling inequality and other social problems head-on. Instead of redistributing wealth, we are promised that growth will fix poverty. Protecting and cleaning up the environment can only be undertaken when they do not compromise growth. Every day we perform roles that serve economic growth: doing wage labor, raising future workers, acquiring marketable skills, pursuing efficiency, competing with one another, ordering stuff online, investing our savings, hoping for more jobs to become available. It took centuries to train people to act like machines for making money for others. Authorities had to take the land people lived from and destroy institutions for meeting needs collectively, in common. Growth relies on exploitation, degrowth scholars argue. To increase total production, capitalists and governments must capture economic surplus and reinvest it to expand productive capacity. To get their hands on society’s surplus, they have to take the fruits of work that women, slaves, communities, and ecosystems do without pay, or remunerate wage laborers less than the full value of what they produce.

17.4

The path to degrowth

Degrowth, accordingly, could reduce exploitation. Degrowth’s champions see themselves as natural allies of the international movement for environmental justice, since shrinking rich economies would reduce outward pressure on the frontiers of extraction. New dams, mines, mega-farms, and factories in the Global South often support consumption in the North. Again, ecological economists might do well to consider it the other way around: reducing exploitation might require or even bring about degrowth. Activists fighting against expanding extraction and pollution are literally blocking growth. The degrowth hypothesis is not “produce less and we will be better off”; it is that if governments and civil society work toward making us better off—ecologically, socially, psychologically, climatologically—then that will entail producing less. So, we need economies that can handle shrinking. Policies for ending nations’ dependance on growth include reducing working hours,

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scaling down harmful sectors like the military, and providing universal access to necessities like housing and health care. Ecological macroeconomists have modeled how, in combination, these policies might allow economies to deliver well-being while contracting. Even so, degrowth advocates hesitate to put forward a blueprint for a post-growth future. Instead, degrowth brings together diverse ideas about, and examples of, non-growing economies. Overgrown societies can learn from Indigenous peoples, peasant societies, ancient civilizations, our grandparents, the poor, and other movements from the tapestry of alternatives. Surveys and experiments find that residents of rich nations are more amenable to degrowth than detractors assume (Drews and Reese 2018; Tomaselli et al. 2021). In the 2018 Yale Climate Opinion poll, a majority of respondents in all 50 United States prioritized environmental protection over economic growth. Yet degrowth’s leading thinkers tend to posit that a transformation beyond growth will come from movements, not votes, even as they acknowledge the need for policy changes to facilitate it. At any rate, degrowth will probably be an outcome of social change toward justice and sustainability, not the idea that propels such change. The outpouring of recent scholarly literature and popular press on degrowth is evidence of the idea’s importance. Numerous books on degrowth have come out since 2018: Degrowth (Kallis 2018), Towards a Political Economy of Degrowth (Chertkovskaya et al. 2019), Exploring Degrowth (Liegey and Nelson 2020), The Case for Degrowth (Kallis et al. 2020), The Degrowth Alternative (Stuart et al. 2020), Less is More: How Degrowth Will Save the World (Hickel 2021), The Future is Degrowth (Schmelzer et al. 2022), as well as deep dives on particular topics such as Housing for Degrowth (Nelson and Schneider 2018) and Food for Degrowth (Nelson and Edwards 2020). Well-known economists’ and pundits’ attacks on degrowth (Milanović 2021; Smith 2021) further illustrate that the debate is gaining momentum. Degrowth, in short, is stopping growth and organizing to live well together without it. In this way, it brings together resistance and alternatives, institutional and extra-institutional politics. Ending growth

justly will require both legislation and direct action. Non-growing societies can meet everybody’s needs through combinations of guaranteed government services, community mutual aid, and self-provisioning. Societies undertaking a degrowth transition will expropriate private riches and construct public wealth. Sam Bliss and Giorgos Kallis

References

BP. 2020. “Statistical Review of World Energy 2020,” 69th ed. https://​www​.bp​.com/​en/​global/​ corporate/​energy​-economics/​statistical​-review​ -of​-world​-energy​.html Brand, Ulrich, Barbara Muraca, Éric Pineault, Marlyne Sahakian, Anke Schaffartzik, Andreas Novy, Christoph Streissler, et al. 2021. “From Planetary to Societal Boundaries: An Argument for Collectively Defined Self-Limitation.” Sustainability: Science, Practice and Policy 17 (1): 265–92. https://​doi​.org/​10​.1080/​15487733​ .2021​.1940754 Chertkovskaya, Ekaterina, Alexander Paulsson, and Stefania Barca, eds. 2019. Towards a Political Economy of Degrowth. London: Rowman & Littlefield. Daly, Herman E. 1973. Toward a Steady-State Economy, illustrated ed. San Francisco: W.H. Freeman. Daly, Herman E. 2014. From Uneconomic Growth to a Steady-State Economy. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Drews, Stefan, and Gerhard Reese. 2018. “‘Degrowth’ vs. Other Types of Growth: Labeling Affects Emotions but Not Attitudes.” Environmental Communication 12 (6): 763–72. https://​doi​.org/​10​.1080/​17524032​.2018​ .1472127 Easterlin, Richard A., Laura Angelescu McVey, Malgorzata Switek, Onnicha Sawangfa, and Jacqueline Smith Zweig. 2010. “The Happiness–Income Paradox Revisited.” Proceedings of the National Academy of Sciences 107 (52): 22463–68. https://​doi​.org/​10​ .1073/​pnas​.1015962107 Georgescu-Roegen, Nicholas. 1971. The Entropy Law and the Economic Process. Cambridge, MA: Harvard University Press. Georgescu-Roegen, Nicholas. 1979. Demain la Décroissance: Entropie, Écologie, Économie [Tomorrow, Degrowth: Entropy – Ecology – Economy]. Translated by Ivo Rens and Jacques Grinevald. Lausanne: Pierre-Marcel Favre. Grubler, Arnulf, Charlie Wilson, Nuno Bento, Benigna Boza-Kiss, Volker Krey, David L. McCollum, Narasimha D. Rao, et al. 2018. “A Low Energy Demand Scenario for

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102  Elgar encyclopedia of ecological economics Meeting the 1.5°C Target and Sustainable Development Goals without Negative Emission Technologies.” Nature Energy 3 (6): 515–27. https://​doi​.org/​10​.1038/​s41560​-018​-0172​-6 Haberl, Helmut, Dominik Wiedenhofer, Doris Virág, Gerald Kalt, Barbara Plank, Paul Brockway, Tomer Fishman, et al. 2020. “A Systematic Review of the Evidence on Decoupling of GDP, Resource Use and GHG Emissions, Part II: Synthesizing the Insights.” Environmental Research Letters 15 (6): 065003. https://​doi​.org/​10​.1088/​1748​-9326/​ab842a Hickel, Jason. 2021. Less Is More: How Degrowth Will Save the World. London: Windmill Books. Illich, Ivan. 1973. Tools for Conviviality. New York: Harper & Row. Kallis, Giorgos. 2018. Degrowth. New York: Agenda Publishing. Kallis, Giorgos, Susan Paulson, Giacomo D’Alisa, and Federico Demaria. 2020. The Case for Degrowth, 1st ed. Medford, MA: Polity. Keynes, John Maynard. 1932. “Economic Possibilities for Our Grandchildren (1930).” In Essays in Persuasion, 358–73. London: Harcourt Brace. Keyßer, Lorenz T., and Manfred Lenzen. 2021. “1.5°C Degrowth Scenarios Suggest the Need for New Mitigation Pathways.” Nature Communications 12 (1): 2676. https://​doi​.org/​ 10​.1038/​s41467​-021​-22884​-9 Krausmann, Fridolin, Karl-Heinz Erb, Simone Gingrich, Helmut Haberl, Alberte Bondeau, Veronika Gaube, Christian Lauk, Christoph Plutzar, and Timothy D. Searchinger. 2013. “Global Human Appropriation of Net Primary Production Doubled in the 20th Century.” Proceedings of the National Academy of Sciences 110 (25): 10324–9. https://​doi​.org/​10​ .1073/​pnas​.1211349110 Latouche, Serge. 2009. Farewell to Growth. Translated by David Macey. Medford, MA: Polity. Liegey, Vincent, and Anitra Nelson. 2020. Exploring Degrowth: A Critical Guide. London: Pluto Press. Meadows, Donella H., Dennis L. Meadows, Jorgen Randers, and William W. Behrens III. 1972. The Limits to Growth. New York: Universe Books. Milanović, Branko. 2021, February 20. “Degrowth: Solving the Impasse by Magical Thinking.” Brave New Europe (blog). https://​ braveneweurope​.com/​branko​-milanovic​

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-degrowth​-solving​-the​-impasse​-by​-magical​ -thinking Mill, John Stuart. 1848. “Of the Stationary State.” In Principles of Political Economy, with Some of Their Applications To Social Philosophy, Book IV, Chapter 6. London: John W. Parker. Nelson, Anitra, and Ferne Edwards, eds. 2020. Food for Degrowth: Perspectives and Practices, 1st ed. Routledge. Nelson, Anitra, and François Schneider, eds. 2018. Housing for Degrowth: Principles, Models, Challenges and Opportunities. London: Routledge. Olson, Mancur. 1973. “Introduction.” Daedalus 102 (4): 1–13. Parrique, Timothée, Jonathan Barth, François Briens, Christian Kerschner, Alejo Kraus-Polk, Anna Kuokkanen, and Joachim H. Spangenberg. 2019. “Decoupling Debunked – Evidence and Arguments against Green Growth as a Sole Strategy for Sustainability.” European Environmental Bureau. https://​eeb​.org/​library/​ decoupling​-debunked/​ Schmelzer, Matthias, Andrea Vetter, and Aaron Vansintjan. 2022. The Future Is Degrowth: A Guide to a World Beyond Capitalism. New York: Verso. Schumacher, Ernst Friedrich. 1973. Small Is Beautiful: Economics as If People Mattered. New York: Harper & Row. Smith, Noah. 2021, September 6. “People Are Realizing That Degrowth Is Bad.” Substack newsletter. Noahpinion (blog). https://​ noahpinion​.substack​.com/​p/​people​-are​ -realizing​-that​-degrowth Stuart, Diana, Ryan Gunderson, and Brian Petersen. 2020. The Degrowth Alternative: A Path to Address Our Environmental Crisis?, 1st ed. New York: Routledge. Tomaselli, Maria Fernanda, Robert Kozak, Robert Gifford, and Stephen R.J. Sheppard. 2021. “Degrowth or Not Degrowth: The Importance of Message Frames for Characterizing the New Economy.” Ecological Economics 183 (May): 106952. https://​doi​.org/​10​.1016/​j​.ecolecon​ .2021​.106952 World Health Organization (WHO). 2020, December 9. “The Top 10 Causes of Death.” https://​www​.who​.int/​news​-room/​fact​-sheets/​ detail/​the​-top​-10​-causes​-of​-death www​ .who​ WHO. 2021. “Air Pollution.” https://​ .int/​westernpacific/​health​-topics/​air​-pollution

18. Deliberative ecological economics Introduction

Deliberation is a group-based process of participation, social exchange, reflection, learning, and meaningful debate (Kenter et al., 2016c). In such group-based processes, participants have the opportunity to reflect upon, form, express, and debate their knowledge, perspectives, values, and beliefs (Kenter, 2016a; McCrum et al., 2009; Orchard-Webb et al., 2016; Spash, 2007). Deliberation is much more frequently recognised as important by ecological economists than by neoclassical economic authors, and ecological economists have applied and developed a range of deliberative and analytical-deliberative methods, such as mediated or participatory modelling and various deliberative monetary, non-monetary, integrated, and mixed-method approaches to economic valuation of the environment. Key reasons why ecological economists advocate deliberation include epistemic considerations, such as uncertainty and complexity of environmental issues, and axiological reasons, including the recognition of value plurality and incommensurability. In this entry, I will first discuss the concept of deliberation. Secondly, I will consider these main ecological economic motivations for deliberation, and thirdly, I will consider deliberative methods.

Deliberation

In essence, deliberation is a process to consider, evaluate, or appraise things. Deliberation can be an individual cognitive-reflective process (e.g. Betsch, 2011), such as someone deliberating over their personal decisions or preferences, or a process of social interaction, such as a group of people seeking to establish a common perspective or outcome. Deliberation can also mean a broader process of decision-making, such as Habermas’ ideal of communicative rationality and communicative action where discussion and sense making can generate new knowledge (McCrum et al., 2009) and enhance democratic decision-making (Orchard-Webb et al., 2016; Ranger et al., 2016). Communicative rationality involves seeking actions that are based on reasoned understanding between

citizens or actors in society. This can be achieved through inter-subjective communication between these actors where they formulate views by reflectively considering the views of others in the communicative process (Zografos and Howarth, 2010). Ecological economists most commonly refer to deliberation as some sort of group process to enhance the elicitation of values and preferences (e.g. Kenter, 2017; Lo, 2013; Spash, 2007; Zografos and Howarth, 2010). This recognises that people’s values and preferences are often only partially formed, and open to change – if they were fully formed and fixed it would be pointless to deliberate them. Consequently, deliberative ecological economists argue that preferences and values need to be generated through some kind of transformative process of deliberation and learning (Kenter et al., 2014b, 2011; Lecture and Norgaard, 2007; Parks and Gowdy, 2013; Spash, 2008). An important aspect of the notion of communicative rationality is that deliberative processes, and communicative ‘actions’ within them, are non-coercive (Dryzek, 2002); it is the force of argument that is decisive, rather than the potential relations of power between participants (Orchard-Webb et al., 2016). This distinguishes deliberative democracy from representative democracy, where decisions are the result of political power expressed as political alliances achieving majorities, though deliberative processes are often undertaken within, and can be constrained by broader representative systems (Grӧnlund et al., 2015). As such, deliberative processes often strive for consensus and emphasise the importance of procedural justice (Zografos and Howarth, 2010). However, consensus views are not always desirable or achievable (Sagoff, 1998), and deliberative processes can also result in the recognition of a diversity of values, achieving outcomes that account for reasonable differences and that can enhance the legitimacy of decision-making processes (Lo, 2013; Ranger et al., 2016). Deliberative ecological economists have engaged both with the political science field of deliberative democracy, and the applied social-environmental field of participation in environmental management. These fields consider deliberation in somewhat different ways. While the former tends to focus on mini-publics (groups of citizens that are carefully sampled and engaged with as a rep-

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resentation of whole populations; Grӧnlund et al., 2015), the latter tends to focus on purposive sampling of groups of stakeholders (Kenter, 2017). Here representation is sought by careful stakeholder analysis to ensure each relevant interest and viewpoint is included (Prell et al., 2009; Reed et al., 2009).

Motivations for deliberation

Ecological economists have argued for deliberation for both epistemic reasons, pertaining to knowledge, and for axiological reasons, pertaining to value theory. In terms of the first, scientific knowledge production in conventional economics approaches, and also in environmental sciences, is typically considered as an expert-led, problem-solving approach, testing structured hypotheses within accepted analytical frameworks (Kuhn, 1962). Many ecological economists have embraced ‘post-normal’ science critiques of this ‘normal’ scientific approach that have argued that it is not sufficient for real-world decisions (Funtowicz and Ravetz, 1993). Environmental problems are marked by high levels of uncertainty that often cannot be reduced, leading to incomplete understandings of problems that may also be contested (Ainscough et al., 2018). Secondly, applied science is driven by decision-making that looks towards evaluating more or less desirable potential futures (Norgaard et al., 2009). This means that knowledge generation is itself constrained by the normative parameters of what is and is not desirable. The simple act of seeking evidence for one thing, rather than another, already involves value judgements, and evidence is frequently used strategically (Hockley, 2014). As discussed in Chapter 59 (this volume), strategic motivations often drive the desire to monetise environmental values. Finally, different traditions across the social and natural sciences are themselves constrained by disciplinary framings that need to be overcome to assess complex interdisciplinary problems (Lecture and Norgaard, 2007) To exemplify these issues, in the UK, management of upland peatland ecosystems has traditionally involved burning mosaics to improve habitats for game birds. Increasing focus on the carbon stock of peatlands has led to calls by many conservationists to end burning practices, based on various scientific studies. However, moorland managers with Jasper Kenter

game interests also justify their competing claims that burning does not significantly damage carbon stocks based on scientific evidence. Banning burning and limiting grazing on upland peat could be detrimental to local economies, but the economic value of carbon stocks, ongoing carbon sequestration, and co-benefits would almost certainly exceed costs (Glenk and Martin-Ortega, 2018; Moxey and Moran, 2014). The debate thus centres around conflicts between precautionary arguments and vested economic interests, as well as the larger-scale non-market benefits versus local costs. The problem of what is the ‘right’ management for peatland can thus not be fully addressed by normal science and environmental economic approaches, but is a matter of value judgement, with different values embedded in the different ways peatlands and their management are framed (Zimmermann et al., 2022). Ecological economists have also argued for deliberation on the basis of value theory. Values include broad, transcendental values, our overarching principles and life goals such as fairness and protecting nature, and more specific, contextual values, the opinions we have about the value or worth of something, such as the cultural value of a woodland (Kenter et al., 2015). Finally, they include value indicators, which measure importance in qualitative or quantitative, monetary, or non-monetary terms (Kenter et al., 2015). In conventional economic analysis, contextual values are considered to be expressed in preferences, reflected in willingness to pay (WTP). In expressing WTP, we are assumed to trade-off our contextual values and all underlying transcendental values. However, ecological economists and environmental philosophers have long questioned that such choices are willing ‘trade-offs’ of something less valuable against something more important, but involve difficult moral sacrifices that cannot be used to infer the value of the things that are ‘traded-off’ (Holland, 2002; Keat, 1997; O’Neill et al., 2008). Indeed, in environmental valuations, participants can struggle or refuse to make such trade-offs (Isacs et al., 2022; Kenter et al., 2011). Different contextual values serve different transcendental ends. In the peatlands example, the importance of carbon and biodiversity relates to our goals of protecting the environment, while burning relates to tradition and local economic opportunities. Ecological econo-

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mists have argued that such ends are fundamentally different, cannot be meaningfully compared in a single metric such as money, and can only be ‘weakly’ compared in deliberative processes through practical judgement (Martínez-Alier et al., 1998). Such issues of incommensurability are also discussed in Chapter 52. A further issue is that, in conventional economic approaches, normative assumptions are made about the way values should be aggregated across individuals (Parks and Gowdy, 2013), albeit often implicit in technical guidance (Hockley, 2014). Methods for valuation and policy appraisal thus have their own ‘meta-values’ (Kenter et al., 2016a) about how values should be considered and aggregated, and these values should be opened to debate. Deliberative ecological economists have had a great interest in both the way that different valuation methods, as ‘value articulating institutions’ (Vatn, 2009) elicit different types of values, and how political institutions privilege and exclude different values with implications for sustainability and equity (Martínez-Alier, 2002; Zografos and Howarth, 2008). Deliberative processes are also appealing for understanding values in that they provide participants with time to think and thus time to reflect upon and form their values (Alvarez-Farizo and Hanley, 2006). Deliberation is however not limited to information, but might also address rights, responsibilities, equity, fairness and other transcendental values, and how to deal with questions of uncertainty (Kenter et al., 2016c). As discussed in more detail in Chapter 59 (this volume), in many real-life cases, non-monetary concerns such as rights can be more salient in decision-making than can monetary values.

Deliberative methodologies

Deliberative methodologies are diverse with roots in different domains of research and practice. They can be loosely organised in deliberative-democratic, analyticaldeliberative, deliberative-interpretive, and deliberative-psychometric methods (Table 18.1), though these categories can overlap or be combined. Deliberative-democratic methods, such as in-depth discussion groups and citizens’ juries, are grounded in political theory. They include diverse techniques

that allow participants to ‘confer, ponder, exchange evidence, reflect on matters of mutual interest, negotiate and attempt to persuade each other’ (Stern and Fineberg, 1996, 73). Through such processes, participants are encouraged to develop and express their views as different perspectives and evidence are considered. Outcomes of deliberative methods are typically qualitative, such as priority lists, recommendations, and verdicts. Analytical-deliberative methods such as deliberative monetary valuation (DMV) and deliberative multicriteria analysis (MCA)1 involve often elaborate approaches integrating deliberative-based techniques with more formal evaluative and decision-support tools. Outcomes typically include quantitative rankings or ratings or monetary values. DMV can be further characterised on a spectrum from deliberative democratic monetary valuation (DDMV) to deliberated preferences (Kenter, 2017; Martino and Kenter, 2023). DDMV seeks to identify aggregate social value or fair prices (e.g. for environmental goods) through valuations based on deliberative theory, while deliberated preferences methods adapt stated preference valuation common to neoclassical environmental economics to elicit informed individual WTP. Most studies to date are situated towards the latter end of the spectrum, building on a lineage of studies originating in ‘market stall’ contingent valuations (Bunse et al., 2015; Lienhoop et al., 2015; Lienhoop and MacMillan, 2007; Schaafsma et al., 2018) The spectrum between DDMV and deliberated preferences reflects the basic challenge for DMV that the languages of political citizenship and communicative reason are fundamentally different to the language of instrumental preferences (see Chapter 59). While this can be conceived of as a creative tension that mirrors the challenges policy makers are often faced with (Bartkowski and Lienhoop, 2018; Kenter, 2016b), how this is resolved tends to depend on the goal of the valuation: ‘economisation’, ‘moralisation’, or ‘democratisation’ (Lo and Spash, 2012). Preference economisation primarily seeks to utilise deliberation to ease the respondent’s cognitive burden associated with expressing stated preference monetary values. Preference moralisation seeks to use deliberation to extend moral concern to the environment. Choice democratisation in their Jasper Kenter

106  Elgar encyclopedia of ecological economics Table 18.1 Examples of deliberative methods Method

Description

Deliberative-

In-depth discussion

Group discussions (often repeated and usually involving 4–8 people) during which

democratic

groups

participants shape the terms of the discussion, develop themes in ways relevant to their own needs and priorities

Citizens’ juries

A small cross-section of the public who come to a considered judgement about a stated policy issue/problem through detailed exposure to, and scrutiny of, the relevant evidence base. The group responds by providing a recommendation or ‘verdict’

Deliberative opinion

A technique designed to observe the evolution of the views of a large citizen test

polls

group as they learn about a topic. Typically, the group votes on the issues before and after an extended debate

Analytical-

Participatory modelling The involvement of stakeholders in the design and content of analytical models

deliberative

(e.g. to represent ecosystems and their benefits under different spatial and temporal conditions) Deliberative monetary

Techniques that use formal methods of group deliberation to come to a decision

valuation

on monetary values for environmental change. Includes deliberated preferences approaches allied to survey-based techniques (contingent valuation or choice experiments) or deliberative democratic monetary valuation approaches that do not use conventional survey or econometric-based techniques to establish values (e.g. by incorporating citizens’ juries)

Deliberative

Techniques that involve groups of stakeholders designing formal criteria (in

multicriteria analysis

monetary or non-monetary quantitative or qualitative terms) against which to

Interpretive-

Participatory mapping/

A group of stakeholders considers or creates a physical or digital map to indicate

deliberative

GIS

landscape features that are valuable (and/or problematic). Participants may also

judge different management options as the basis for making a decision

rate or rank these features for importance. Map layers can also incorporate photo, video, artwork, poetry, and so on Storytelling

Participants are asked to tell stories about their experiences of or in relation to places. These may be reflected upon in a group setting to discuss values related to these experiences

Group interviews

Participants are interviewed about their values, beliefs, and preferences. Group interviews allow for deliberation and are similar to in-depth discussion groups. However, in group interviews, the terms are set by the interviewer rather than the group

Psychometric-

Values compass

deliberative

This method asks participants to consider which of their individual transcendental values are most important by ranking or rating them, and then asks them to discuss the degree to which these values are important for the community, culture, or society. Values can also be ranked or rated on a group basis

Q-method

In this method, participants engage with a range of statements that are sorted according to their agreement, and subsequently statistically analysed. It can be linked to deliberative processes to help structure discussion and assess and organise different viewpoints, or to analyse how deliberation may shift perspectives

Source: Adapted with modifications from Fish et al. (2011), Kenter (2016c), and Kenter et al. (2014a).

view moves beyond Habermasian consensus to emphasising value plurality. Rather than following standardised procedures, such Jasper Kenter

an approach centres on key principles and requirements in relation to process. Deliberative MCA, also known as social multicriteria evaluation, has long been pro-

Deliberative ecological economics  107

moted as a superior alternative to neoclassical monetary valuation methods by ecological economists because of its ability to explicitly structure problems according to multiple values (Martínez-Alier et al., 1998). In MCA, different alternatives (e.g. policy options) are considered and scored against multiple criteria, such as economic benefits and costs; broader social, ecological, and cultural values and concerns; and technical criteria like capacity for implementation or likelihood of policy compliance. While this provides a more pluralistic approach than using money criteria alone, MCA is challenged by similar issues to monetary valuation in terms of how to aggregate across different criteria and across individuals (Murphy et al., 2017). Deliberative processes can thus be used to bridge criteria and try to build consensus on desirable outcomes (Ranger et al., 2016), which can lead to different rankings of alternatives than aggregation of individual preferences (Murphy et al., 2017). Sophisticated MCA tools have been developed congruently with ecological economic premises about complexity, value plurality, and incommensurability, such as by including imprecise assessments, and by allowing options for partial compensation and incomparability between criteria (Munda, 2004). Participation can be further enhanced by empowering participants to select or define options and criteria themselves, though again this can be constrained by the decision context (Proctor and Drechsler, 2006). Ecological economists are also increasingly combining deliberative methods with psychological approaches, such as Q methodology (Isacs et al., 2022), and interpretive approaches, such as ethnography and storytelling (Kenter et al., 2016b; Ranger et al., 2016). This reflects the cultural and relational turn in environmental assessments with more emphasis on complex, often place-based connections between people and nature (Fish et al., 2016; Himes and Muraca, 2018) and the shared nature of such values (Cumming and Norwood, 2012; Ishihara, 2018). Sometimes complex methodologies and elaborate multistage designs are developed that integrate different deliberative democratic, psychological, interpretive, and analytical tools. Such customised integrated deliberative valuation approaches are relatively resource intensive but have the greatest potential to

bridge multiple values and valuation languages. Co-design by decision-makers also allows such customised approaches to directly interface with decision context using post-normal scientific approaches (Kenter, 2016a; Orchard-Webb et al., 2016; Ranger et al., 2016; Slater et al., 2020). Regardless of which group deliberative approach is applied, a key pitfall of deliberative processes relates to who is represented around the table. This relates to both the inclusiveness and authenticity of the deliberation. Inclusiveness pertains to the effective representation of relevant stakeholders (in purposive sampling-based approaches, such as stakeholder workshops) or social and cultural groups (in approaches based on ‘mini-publics’, such as citizen juries), but also the degree to which all relevant discourses receive attention (Orchard-Webb et al., 2016; Schouten et al., 2012). Authenticity pertains to the degree to which deliberation reflects genuine, non-coercive communicative processes (Schouten et al., 2012). It is hard to ignore the often unequal social relations and institutions outside of the valuation setting, which will influence participants to voice their opinions and concerns (O’Neill, 2013). Differences that can lead to power dynamics can include social status, political influence, class, education, and experience with deliberative processes. Managing such dynamics demands both professional facilitation and explicit consideration in deliberative process design (Kenter et al., 2016c). Fortunately, ecological economists can build on extensive experience of facilitation and design across deliberative fields, from deliberative democracy to action research and stakeholder participation in environmental management. A further question relating to the democratic capacity of deliberative processes and methods regards their consequentiality (Dryzek, 2009). Consequential deliberation means that discourses and values arising from the deliberation are reflected in institutional outputs and outcomes, such as policy, legislation, or voluntary standards. Deliberation is often framed in terms of participation in decision-making processes, but deliberative discourses can struggle to compete with prevailing institutional aspects. For example, consequentiality can be hampered in multi-stakeholder market-based initiatives, such as roundtables, due to prevalent market Jasper Kenter

108  Elgar encyclopedia of ecological economics

relations (Schouten et al., 2012). As such, the transformative potential of deliberation may be limited unless deliberative processes are also empowered to challenge such structural issues.

Future directions

While ecological economists have engaged with deliberative theory for decades, deliberative studies haven’t burgeoned, particularly in terms of genuinely deliberative democratic approaches, and post-normal science approaches that integrate deliberation and co-production of knowledge with communities of stakeholders. Considering publications within journals such as Ecological Economics, deliberative ecological economics still appears to make up a small subset within the broader school. This may be related to the challenge of more fully departing from conventional economic value theory on utilitarian preferences, but also more general challenges of transdisciplinarity and bridging science and practice, such as building capacity within research teams for combining and bridging economic, deliberative, and interpretive methods, brokering their diverging epistemic underpinnings, engaging practical skills such as group facilitation, aligning the different working practices and time horizons of science and policy making, navigating potential tensions between agendas of policy makers and academic independence and freedom, and (under)recognition of applied and policy work in academic institutions, such as in promotion procedures. While attempts have been made to develop deliberative value theory (Bartkowski and Lienhoop, 2018; Howarth and Wilson, 2006; Irvine et al., 2016; Kenter et al., 2019, 2016c; Lienhoop et al., 2015; Orchard-Webb et al., 2016; Spash, 2008), there is still a struggle to align questions of resource efficiency with overarching ecological economic principles of sustainable scale, equity and resilience, and other values. With regard to DMV in particular, technical guidance is mostly geared towards deliberated preferences (Schaafsma et al., 2018) with less consideration about how to facilitate deliberative democratic value formation (Kenter et al., 2016c), and what are acceptable forms and languages of value expression, for example, in terms of the role of emotion (Chapter 59). Qualitative analysis approaches to explicitly evaluate the Jasper Kenter

authenticity of deliberation have been developed (Steenbergen et al., 2003), but have to our knowledge not yet been applied by ecological economists. An interesting avenue of research is also how different languages of valuation and ways of framing value and human–nature relations could represent the more-than-human and support the inclusion of other species within deliberation, also building on fields such as more-than-human geography (Bastian, 2016; Kenter and O’Connor, 2022). In this area, and more broadly, there is also a vast opportunity to engage with diverse interdisciplinary perspectives on value expression and value conflict, for example, from the arts (Edwards et al., 2016). However, this also raises questions around the institutional capacity and demand to engage with more diverse value-knowledge within current decision-making institutions constrained by dominant ideologies of instrumentalism, markets, efficiency, and growth. A key hook here is that failures of dominant thinking to recognise shared values can lead to embarrassing policy failures (Kenter et al., 2014a), whereas deliberative processes can improve the perceived legitimacy and compliance with policies (Ranger et al., 2016), which is highly salient to authorities with often limited resources for policy implementation and enforcement. Further research may consider explicitly what factors and conditions hamper and support post-normal deliberative processes (e.g. in terms of their impact on policy, their perceived legitimacy by policy makers and participants, their efficiency in terms of resources/cost); whether they can improve outcomes in terms of sustainability, equity, and social-ecological resilience; and how they can be used strategically to challenge old economic and mainstream new ecological economic institutions. Jasper Kenter

Note 1.

This is also referred to as deliberative or participatory multi-criteria evaluation, or multi-criteria decision analysis/evaluation.

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19. Discounting and climate change The rationale for discounting

Discounting allows us to compute the present value of financial flows that will take place in the future. Discounting is needed in benefit–cost analysis to calculate net present values – the key criterion for investments. At a more global level, discount rates relate to investment rates: the lower the former, the higher the latter. As such, discounting reflects the balance between present and future well-being. As Irving Fisher (1930) established, discounting reflects both the productive nature of our economies and an individual’s or society’s impatience. In a world without market failure, tax, and risk, the return on investment would be equal to the social rate of time preference, which accordingly is the sum of the pure rate of time preference and the product of the growth rate of per capita income multiplied by the elasticity of the marginal utility of income (Ramsey, 1928), sometimes named the ‘wealth effect’. We generally discount future amounts of money using a discount rate that is constant through time, leading to ‘exponential discounting’. As a result, values in the far-distant future are reduced to very low levels. For example, damages of €1 million in 100 years have a present value of €52 000 at a discount rate of 3 per cent (annually). At a discount rate of 8 per cent, their present value is only €455, and the present value of the sum of an infinite series of discounted annual amounts of €1 equals €12.5. While the first 40 years account for more than €12; values beyond this point are negligible.

The dispute over the Stern Review

One argument often made is that discounting is ‘unethical’: people’s welfare should not be valued less simply because they live at a different time. Pure time preference would be acceptable as far as it reflects individuals’ choices – but not in an intergenerational context. In his influential review of the economics of climate change, Nicholas Stern (2006) put the rate of pure time preference at 0.1 (to account for the possibility of extinction of the

human species), and the elasticity of income at 1. Assuming a per capita growth rate of 1.3 per cent, Stern gets a discount rate of 1.4 per cent, leading to high present values of future climate change damages, and therefore justifying strong climate change mitigation action. The economists who disagreed with the conclusions of the Stern Review focused their criticism on its low rate of discount. For example, Nordhaus (2007) noted that this number does not match the observed market rates of interest. Defenders of the conclusions of the Stern Review often did not support Stern’s arguments relative to the discount rate but instead referred to the uncertainty on the future states of the word, and to the relative evolution of prices resulting from the non-substitutability of natural assets. These are the main points discussed here. As Weitzman (2007, p. 703) expressed it in a comment on the Stern Review, ‘spending money to slow global warming should perhaps not be conceptualised primarily as being about consumption smoothing as much as being about how much insurance to buy to offset the small chance of a ruinous catastrophe that is difficult to compensate by ordinary savings’. Today this ‘chance’ does not appear to be that small – but this only increases the relevance of Weitzman’s vision.

Discounting is not unfair to future generations

Why is Stern’s argument on discounting not fully convincing? Setting the pure time preference at or very close to 0 on ethical grounds questions the other component of the discount rate, the ‘wealth effect’. If future generations are richer than the current one, there is little justification of depriving additional money from the current, relatively poor generation to increase wealth of subsequent ones. In other words, if one chooses to be ethically prescriptive on pure time preference and set it at or near 0, consistency requires us to use a similar approach, but with opposite results, in relation to the wealth effect. Discounting the future does not appear unethical, for if discounting the utility of future generations might be, discounting their consumption might not be, provided per capita economic growth is real. As Baumol (1968, p. 800) wrote,

112

Discounting and climate change  113 a redistribution to provide more for the future may be described as a Robin Hood activity stood on his head – it takes from the poor to give to the rich. Average real per capita income a century hence is likely to be a sizeable multiple of its present value. Why should I give up part of my income to help support someone else with an income several times my own?

In this sense, an ethical appraisal of discounting does not conflict with Fisher’s lesson: the productive nature of the economy legitimates discounting. It is possible, however, that people receiving future benefits are not better off than those incurring current costs. For example, this might apply in the case of climate change; those more likely to reduce greenhouse gas emissions today are people in industrialised countries, while those more likely to benefit from reduced emissions in the future are the poor in developing countries lacking resources for adapting to climate change. Given the extent of disparity between developed and developing counties, people from developing countries in the future may well still be poorer than current people in developed countries. However, does this mean that in case of climate change one should use a zero or even negative discount rate, as some have argued? Probably not. Funds spent in climate change mitigation have opportunity costs. It may be more efficient to devote the resources to development projects to help people in developing countries to achieve faster economic development. Climate change mitigation investments should thus compete with other development projects, using discount rates that are appropriate for projects in developing countries. Given the scarcity of capital, these are usually higher, not lower, than rates used in developed countries. Discounting per se is not unfair, provided future generations are effectively richer. Indeed, discounting helps to ensure the greater wealth of future generations by allowing them to select efficient investments. Discounting may also have an environmental upside, for in its absence, or if we used discount rates that are too low, many more investments would be warranted, which would increase the pressure on natural resources and ecosystems with little additional benefit for the populations.

The paradoxes of long-term discounting

Rabl (1996) points out a difficulty in the use of standard discount rates over the very long term: the rate of return on marginal investment cannot be durably higher than the growth rate of the economy. This would lead to paradoxes: any investment, however small, but with a return rate greater than the growth rate of the economy would have, after enough time has elapsed, an output greater than the whole economy: clearly an absurdity. Over long periods of time, compound interest rates give dramatic results. One gram of gold saved with an interest rate of 3.25 per cent when Jesus was born would be worth today 6000 billion tonnes of gold – the weight of planet Earth. This does not mean that marginal rates of return on investments cannot be higher, at any time, than the growth rate of the economy; part of the explanation for this is that the output of these investments is largely consumed, and only in part reinvested. Benefit–cost analysis supposes that possible beneficiaries of the investment or policy under scrutiny could, in principle, compensate any losers. Discounting future damages (e.g. resulting from climate change) that could be avoided thanks to some investment (e.g. emissions mitigation) rests on the implicit hypothesis that alternative investments would have a rate of return at least equal to the discount rate used. However, rates of return higher than gross domestic product (GDP) growth rates cannot be sustained for ever. Thus, discount rates in the long run must come close to the growth rate of the economy. Rabl suggests a two-tier discounting procedure, using the conventional rate for a short period (30 years, for example) and then a reduced rate for intergenerational effects, equal to the rate of long-term economic growth. One problem with that proposal is time inconsistency, as Solow (1999) notes after Ramsey (1928). Using Rabl’s suggested approach, the value of a unit of capital in 2030, equal to the discounted sum of its future net benefits, will differ depending on whether it is calculated in 2000 or in 2030.

Discounting uncertain futures

Can we be sure of future growth rates of per capita welfare, and can we be sure that Cédric Philibert

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the environmental damages we are currently creating will not harm future growth? Of course not. The dramatic collapse of biodiversity and climate change strongly suggest the opposite. Hence the discount rate to give a present value of future environmental damages, and thus determine the proper level of investment in mitigating them, cannot be set exogenously. Let us imagine two states of the economy 100 years hence. One corresponds to slow growth (for which a low discount rate is appropriate), the other to high growth (leading to a high discount rate). Let us consider them as equally probable. Let us now consider the present value of a sum of money from 100 years hence. Using the standard approach of decision theory, it should be the weighted average of the net present values computed using the two discount rates. However, as noted by Weitzman (1998), this average is dominated by the value computed using the low discount rate. In the high discount rate scenario, the present value is discounted to a trivially small level. As a result, if future growth is uncertain, the discount rate should come progressively closer to the ‘lowest possible’ discount rate. A risk-averse attitude would further stress this argument. Newell and Pizer (2003) brought this insight to their study of uncertain discount rates. Their starting point was rates of return on investments based on observed risk-free market rates. Over long periods of time they computed yearly benefits accruing from climate change mitigation. Results obtained using uncertain discount rates were compared with results obtained using a fixed discount rate set at the expected value of the uncertain distribution. Because unexpectedly low discount rates raise valuations by a much larger amount than unexpectedly high discount rates reduce them, the uncertainty about the discount rate always raises the valuation of future benefits. Newell and Pizer (2003) concluded that effective discount rates should progressively decline. Using declining discount rates because of uncertainty would not be time inconsistent, although the value of a given unit of capital in 2030 as computed in 2000 may take a lower value in 2030. This value may legitimately change with the passage of time, for the latter progressively reduces the uncertainty on future growth rates (Philibert, Cédric Philibert

1999). In other words, behaviour that would be time-inconsistent in a deterministic world is legitimate state-contingent behaviour in a world with uncertain discount rates (Newell and Pizer, 2003). An expert panel gathered in September 2011 by Resources for the Future revealed some consensus around this conception that included critics of the Stern Review such as Richard Tol and William Nordhaus together with supporters such as Kenneth Arrow, Christian Gollier, Robert Pindyck, Thomas Sterner, Martin Weitzman, and others (Arrow et al., 2012).

Relative prices and discounting

As Krutilla (1967, p. 783) wrote, ‘natural environments will represent irreplaceable assets of appreciating value with the passage of time’. How should this value grow over time? Referring implicitly to the Hotelling (1931) rule regarding the optimal use of non-renewable natural resources, Boiteux (1976, p.  830) writes that, ‘all economic models show that in a growing economy the prices of resources available in strictly limited quantities should be assumed to grow at an annual rate that is at least equal to the discount rate’. As a result, ‘in the long run, the discounting process clears everything that is of secondary importance because it can be controlled by human proficiency, to stress what is essential: i.e., whatever is intrinsically scarce and cannot be reproduced’ (Boiteux, 1976, p. 831). In other words, if correctly valued (given values growing over time), the natural environment will not be disadvantaged by discounting because discounting progressively erases the values of the fruits of one’s labour, but not the irreplaceable environmental assets. However, giving any environmental asset a value growing over time at the pace of the discount rate eventually leads to the paradox discussed by Rabl: over time, this asset will be valued more highly than the rest of the economy. One consequence is that the destruction of an environmental asset (e.g. extinction of a species) would have the same present cost whenever it happens. And delaying damages would have no value. However, delaying irreversible damages leaves open the possibility that it will not happen due to technical progress or other developments.

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The current collapse of biodiversity and wildlife – sometimes termed the ‘sixth extinction’ – is now recognised as a major environmental challenge; however, species do not last forever, and the evolution has not begun with industry. In view of this, environmental assets that are neither reproducible nor substitutable should be given a value growing over time at a rate close to, but slightly less than, the discount rate. As a result, environmental assets would be submitted to what Fisher and Krutilla (1975) called ‘effective discounting’, but at a very low rate, which we might call ‘slow effective discounting’. The lack of effective discounting would give the current generation an unlimited responsibility with respect to future generations. As argued by Ricoeur (1995, p. 68), Completely ignoring the side effects of the action would make it dishonest, but unlimited responsibility would make it impossible. It is indeed a sign of human limitations that the disparity between the desired effects and the innumerable consequences of the action is itself unmanageable and calls upon the practical wisdom gained throughout the history of earlier trade-offs. A happy medium must be found between escaping from the responsibility for consequences and the inflation of infinite responsibility.

Other analysts have also underlined the evolution of relative prices. For Neumayer (1999), discounting is not the issue – substitutability is. Valuing environmental assets in monetary terms rests on the assumption that environmental and other values are substitutable for each other. Hoel and Sterner (2007) analyse a conceptual model of the economy consisting of one (conventional) sector which grows ‘forever’ and another sector (say, ‘environmental services’) that is constant (or maybe even declining due to pollution). The environmental sector can see its share of the economy grow in value terms, despite becoming physically smaller in comparison to the growing sector due to rising relative prices. Sterner and Persson (2007) illustrate the implications in the case of climate change. They show that an emission scenario in a case with a high discount rate but in which the increasing relative price of the ‘nonmarket’ goods is taken into account is rather close to that of the Stern Review.

The uniqueness of discount rates

A single discount rate for all projects makes sense if they are ‘small’, having no influence on the broad economy: ‘the government is able to pool risks’ (Arrow and Lind, 1970). Building on the reflexions relative to the uncertainty on economic growth and the impact of projects on that growth, as well as risk aversion, more recent developments take account of the systemic macroeconomic risk, denoted φ (phi), and of the elasticity of the future benefits of a specific project on the per capita GDP, denoted β (beta). In short, the idea is to better valorise projects that induce more resilience to shocks and penalise more projects that increase economic risks. In France, for example, Emile Quinet (2013) recommended to the French authorities a discount rate, ρ = rf + φ.β, where rf) is the riskless discount rate. He also recommended to take rf = 2.5 per cent and φ = 2 per cent, so that for a project not entailing a deviation of economic growth (β = 1), the discount rate would be 4.5 per cent. A panel under the chairmanship of Roger Guesnerie (2021) has revised this recommendation and proposed, for the period 2021 to 2070, rf = 1.2 and φ = 2 per cent. When β is unknown, it should be considered equal to unity, in which case ρ = 3.2 per cent. While the riskless reference rate has been reduced considering low real interest rates and reduced long-term growth potential of the French economy, the risk premium is unchanged. In practice, Guesnerie, following Quinet, recognises that assessing β project by project can be time and resource consuming, and the idea is instead to use specific βs for project categories. The only βs known today are relative to public projects in transports (1.1 for urban commuters; 1.4 for regional commuters; 1.7 for long-distance travellers; 1.4 for rail freight), and a working group has been tasked to provide values for all the economic sectors. However, it is not entirely clear if and how these refinements will effectively allow project evaluation to better take into account the specifics of individual projects in a given sector, which can have very different long-term consequences, notably for the climate and, more globally, the environment.

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Conclusion

As far as the environment is concerned, the most important point put forward here is that environmental assets that are neither substitutable nor reproducible should be given a value growing over time at a pace close to, but slightly less than, the discount rate. This would result in greater net present values for prevention of future environmental damage and, for example, may justify greater greenhouse gas mitigation efforts in the short term. This reinforces the argument for declining discount real rates based on the uncertainty relative to future growth. Future environmental damages may, in this framework, become so large that they would likely shrink future welfare. The proposal to grow the valuation of environmental assets over time has another important implication: assessment of the long-term consequences of current policies will likely be dominated by environmental values. But environmental assets are only marginally present on current markets, and thus, their monetary value is often hard to estimate. As a result, the present value of future environmental damage increases, but the uncertainty surrounding its estimation increases. This is a clear limit of the cost–benefit framework in which the discounting procedure is essential. However, as Weitzman (2009, p. 18) concluded his examination of the economics of catastrophic climate change, ‘acknowledging more openly the incredible magnitude of the deep structural uncertainties that are involved in climate-change analysis might go a long way toward elevating the level of public discourse about what to do about global warming’. Cédric Philibert

References

Arrow, K.J., and R.C. Lind (1970), Uncertainty and the evaluation of public investment decisions. American Economic Review 60: 364–78. Arrow, K.J., M.L. Cropper, C. Gollier, B. Groom, G.M. Heal, R.G. Newell, W.D. Nordhaus, et al. (2012, December), How Should Benefits and Costs Be Discounted in an Intergenerational Context? The Views of an Expert Panel.

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Resources for the Future, Discussion Paper 12-53, Washington, D.C. Baumol, W.J. (1968), On the social rate of discount. American Economic Review 58: 788–802. Boiteux, M. (1976), A propos de la ‘Critique de la théorie de l’actualisation’. Revue d’économie Politique 5: 828–31. Fisher, A.C., and J.V. Krutilla (1975), Resource conservation, environmental preservation, and the rate of discount. Quarterly Journal of Economics 89: 358–70. Fisher, I. (1930), The Theory of Interest. New York: Kelley and Millman. Guesnerie, R. (Dir.) (2021), Révision du taux d’actualisaation. Paris : France Stratégie. Hoel, M., and T. Sterner (2007), Discounting and relative prices. Climatic Change 84: 265–280. Hotelling, H. (1931), The economics of exhaustible resources. Journal of Political Economy 39: 137–75. Krutilla, J.V. (1967), Conservation reconsidered. American Economic Review 57: 777–86. Neumayer, E. (1999), Global warming: discounting is not the issue, but substitutability is. Energy Policy 27: 33–43. Newell, R.G., and W.A. Pizer (2003), Discounting the distant future: how much do uncertain rates increase valuations? Journal of Environmental Economics and Management 46: 52–71. Nordhaus, W. D. (2007), The Stern review on the economics of climate change. Journal of Economic Literature 45: 685–702. Philibert, C. (1999), The economics of climate change and the theory of discounting. Energy Policy 27: 913–29. Quinet, É. (2013), L’évaluation socioéconomique des investissements publics. Paris: Commissariat Général à la Stratégie et à la Prospective. Rabl, A. (1996), Discounting of long-term costs: what would future generations prefer us to do? Ecological Economics 17: 137–45. Ramsey, F. (1928), A mathematical theory of saving. Economic Journal 38: 543–59. Ricoeur, P. (1995), Le juste. Paris: Éditions Esprit. Solow, R.M. (1999), ‘Foreword’, in Portney, P.R., and J.P. Weyant (eds), Discounting and Intergenerational Equity, pp. vii–ix. Washington D.C.: Resources for the Future. Stern, N. (2006, October 30), The Stern Review Report: The Economics of Climate Change. London: HM Treasury. Sterner, T., and U.M. Persson (2007, July), An even Sterner review. Resources for the Future, Discussion Paper 07-37, Washington, D.C. Weitzman, M.L. (1998), Why the far-distant future should be discounted at its lowest possible

Discounting and climate change  117 rate. Journal of Environmental Economics and Management 36: 201–08. Weitzman, M.L. (2007), A review of The Stern Review on the Economics of Climate Change. Journal of Economic Literature 45: 703–24.

Weitzman, M.L. (2009), On modelling and interpreting the economics of catastrophic climate change. The Review of Economics and Statistics 91 (1): 1–19.

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20. Ecofeminisms This chapter introduces ecofeminist critical thinking and demonstrates how it relates to ecological economics. While decidedly acknowledging that “feminism and environmentalism were both practices first and theory second” (van den Berg, 2019: 55) and thus being critically aware of practical and political contributions by peasant, community, and Indigenous feminist groups, this chapter focuses on theoretical contributions and academic debates. It gives a brief genealogy of ecofeminist scholarship (section 20.1); discusses the questions of essentialism (section 20.2) and dualisms (section 20.3); introduces some theoretical debates and activist struggles in the field of feminisms, decoloniality, and the environment (section 20.4); and links the discussions back to (feminist) ecological economics (section 20.5).

20.1 Ecofeminisms – a brief genealogy

The term “ecofeminism” emerged in the 1970s and describes both a critical theoretical perspective and applied praxis that combines feminist and ecological concerns. Ecofeminists discuss, for example, how gender(ed) relations and power asymmetries have shaped and continue to shape societal relationships with nature and vice versa (Warren, 1996) or how nature exploitation not only relates to patriarchy, but also to racism and coloniality (Shiva & Mies, 1993; Segato, 2014). The most widely received version of academic ecofeminism is rooted in the Anglo-American tradition. And indeed, landmark publications by early ecofeminists like Rachel Carson (1962), Carolyn Merchant (1980), and Val Plumwood (1993), as well as events like the 1980 conference Women and Life on Earth: Ecofeminism in the Eighties in Amherst (US) were important milestones for the development of ecofeminist scholarship. While decidedly acknowledging these contributions, this chapter aims at venturing beyond a reading of ecofeminism as only or mainly rooted in Anglo-American thought. Rather than relegating regional and topical varieties of ecofeminism to the role of case studies, a comprehensive account of ecofeminism should also consult, for example, Latin American ecofeminisms (Gebara, 2003;

Svampa, 2015; Schild, 2019), Indian feminist environmentalism (Shiva, 1988; Agarwal, 1992; Singh, 2013), the German subsistence approach (Mies, 1986; von Werlhof, 1988; Bennholdt-Thomsen & Mies, 2000), and ecofeminist contributions from the African continent (Maathai, 1985; The WoMin Collective, 2017; Chemhuru, 2019), as well as the non-linear and multi-directional flows between these ideas (Lal et al., 2010). While this ambition necessarily remains fragmentary and incomplete, in using the plural form of ecofeminism (ecofeminisms)1, I hope to spark the reader’s interest in the broad variety, entangled histories, and common outlooks of ecofeminist critical thinking around the world.

20.2 Ecofeminisms and the Question of Essentialism

From the 1990s onward, and related to the “discursive turn” in social science, the most prevalent prejudice that is invoked against ecofeminisms – and thus the first issue I hope to clarify in this chapter – is the claim that ecofeminisms, by proposing an assumed biological proximity of women and nature, fall prey to essentialism. For example, in her influential critique of ecofeminism, Janet Biehl (1991: 12) holds that “[u]nlike other feminists, who tried to demolish gender stereotypes as insufferably constraining to women’s full development as full human beings, such ecofeminists enthusiastically begin to embrace some of these same psycho-biological stereotypes.” Engaging with ecofeminist scholarship, however, it becomes evident that the strand of “cultural ecofeminism” that Biehl and other critics refer to is marginal in ecofeminist scholarship.2 Quite to the contrary, a closer reading of early landmark publications (e.g., Merchant, 1980; Mies, 1986; Salleh, 1997) shows that the arguments are decidedly materialist, emphasizing the structural similarity of how (the unpaid care work mostly carried out by) women and nature are exploited and invisibilized by an economic system that, at the same time, fundamentally depends on these spheres. Stemming from the lived experience of performing the lion’s share of care work and of being socialized into this gendered division of labor from a very young age arises what Salleh (2009) calls an “embod-

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ied materialism.” This form of materialism does not start from the assumption of “unchanging ‘essential’ characteristics of sex, gender or nature” (Merchant, 1980: xvi) but from a socially constructed (and yet to be deconstructed) gendered division of labor. Sherilyn MacGregor (2021: 43) emphasizes: “Contrary to how ecofeminism is commonly caricatured, most of those who have engaged with this perspective over the years have theorized connections between nature and gender through labour (not biology or spirit) and the ways in which labour brings people into close contact with the biophysical environment.”

20.3

Ecofeminisms and the question of dualisms

A recurring theme in ecofeminisms is the critique of dualisms, which are integral to a Western epistemology and ontology based on anthropo-, andro-, and Eurocentrism. Going back to 16th-century Enlightenment philosophers like René Descartes and Francis Bacon, who are often regarded as the “founding fathers” of modern science, we find passages like the following: There is therefore much ground for hoping that there are still laid up in the womb of nature many secrets of excellent use. … Neither ought a man to make scruple of entering and penetrating into the holes and corners [of the secrets of nature], when the inquisition of truth is his sole object. (Spedding et al., 1858/2011: 100, 296)

In line with Descartes’ famous mind/body split (cogito ergo sum), which was the first Cartesian dualism, the Bacon quote entails important dualisms where reason (“truth”) and men are seen as ontologically separate from nature and women (“womb”), with the latter being conceptualized as inferior, a domain to be “penetrated” and mastered. Cartesian dualisms and, more generally, Enlightenment ontologies and epistemologies have been problematized by ecofeminist (e.g., Merchant, 1980; Gaard, 1997; von Winterfeld, 2006) and postcolonial scholars alike (e.g., Grosfoguel, 2011; Segato, 2014; Patel and Moore, 2017). Early on, Val Plumwood (1993) showed that hierarchical dualisms lie at the core of the devaluation, exploitation, and destruction of female-codified care work and nature. Building on that, Johana Oksala (2018: 220)

emphasizes that the “feminization of nature and the naturalization of women do not function merely as ideological justifications for an abstract and general logic of domination, but concretely structure the capitalist society through gendered social and economic practices and divisions of labor.” Interventions by post- and decolonial feminists have elaborated on the link between patriarchy and coloniality/racism, showing that the focus on dualisms must not invisibilize hierarchies within categories. For example, Maria Lugones (2010: 757) argues that each “category is characterized in terms of the superior member of the dichotomy,” where “women” are seen as “white women.” Against this background, an intersectional3 ecofeminist critique must be crucially aware of interlocking systems of oppression. Binary thought structures, such as human/ nature, reason/emotion, self/other, male/ female, production/reproduction, heterosexual/queer, white/nonwhite, and colonizer/ colonized, persist until this very day and shape, for example, the way we envision societal relationships with nature. For example, ecological economics emphasizes that the economy is not a separate sphere but fundamentally embedded in society and the biosphere. This is a key difference and crucial step forward when taking mainstream (environmental) economics as a reference and starting point. However, starting from – for example – the Indigenous cosmovision of buen vivir in Ecuador (Acosta & Martínez, 2009) or sentipensar (thinking-feeling) in Colombia (Escobar, 2016), which are based on the ontological inseparability of humans and nature, one cannot help but realize how deeply ingrained the human–nature split is in the Western understanding of what the world is like (ontology) in the first place (Walsh, 2010; Vega Ugalde, 2015). Similarly, debates about whether the concept of “care” should be extended to “caring for nature” do not make sense when starting from non-anthropocentric political ontologies built around interdependence that very naturally include care for nature simply as “care” as there is no subject/object, caregiver/carereceiver split to begin with (Singh, 2013; Quiroga Díaz, 2015; Chemhuru, 2019). In its inherently anti-dualist aspirations, ecofeminism can draw plenty of inspiration from these examples, as well as all kinds of theory and practice that destabilize Corinna Dengler

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either–or hierarchical dualisms and prefigure both–and alternatives, such as queer ecologies (Gaard, 1997; Mortimer-Sandilands & Erickson, 2010; Seymour, 2020).

20.4

Current debates on feminisms, decoloniality, and the environment

Ecofeminisms around the world have put forward the analysis that capitalism is based on the exploitation of what Rosa Luxemburg (1913/2003) has termed non-capitalist milieus. The image of an iceberg, introduced by Maria Mies (1986) and later by J.K. Gibson-Graham (1996), is a useful metaphor to think about this connection:4 “[I]n the course of the last four or five centuries women, nature and colonies were externalized, declared to be outside civilized society, pushed down, and thus made invisible as the underwater part of an iceberg is invisible, yet constitute the base of the whole” (Mies, 1986: 77). Apart from this structural argument that helps us to gain a deeper understanding of the patriarchal, colonial, and ecologically destructive roots of the capitalist growth paradigm, ecofeminisms have always also engaged in prefiguring alternatives. Concepts like the “subsistence perspective” (Bennholdt-Thomsen & Mies, 2000), “earthcare” (Merchant, 1986), or the “sustainability of life” (Pérez Orozco, 2014) start from bottom-up experiences of actually performing everyday “reproductive and earthcare labour” (Barca, 2020: 7), constantly reminding us that ecofeminisms are not only theory, but also critical applied practice. In line with this, ecofeminist activism has always been a core element of ecofeminisms – thereby transgressing theory/practice and science/activism dualisms. For example, early ecofeminist scholarship has been crucially inspired by the 1973 Chipko Movement, which by means of civil disobedience (in the form of tree-hugging) resisted deforestation in Uttarakhand, India (Shiva, 1988), as well as by the 1977 Green Belt Movement in Kenya (Maathai, 1985), where mainly rural women engaged in grassroot reforestation and the (re-)construction of communal knowledge to protect local biodiversity. Until today, ecofeminist activism resists patriarchal, colonial, and extractive capitalism and envisions a transformation to a regeneraCorinna Dengler

tive economy that prioritizes people and the planet and, hence, life over profit. A current example is the ecofeminist pan-African network African Women Unite Against Destructive Resource Extraction (WoMin), which has started building post-extractivist and ecofeminist alternatives to development, taking the position of rural, working-class women in Africa as a starting point for transformation (The WoMin Collective, 2017). The collective’s elaborations on free, prior, and informed consent of communities are important groundwork for struggles against extractivism around the world (Deonandan et al., 2017; Brock & Dunlap, 2018; Fernandes, 2018). Another important debate in the field is that on collective forms of caring for the human and more-than-human (Singh, 2013; Dengler & Lang, 2022). This includes resistance against the enclosure of the commons and their (re-)creation, with commons defined as a historically ubiquitous, non-commodified mode of social and environmental reproduction that is both “an already present reality, especially in the form of existing communitarian forms of social organization, and a perspective anticipating in an embryonic way a world beyond capitalism” (Federici, 2019: 4). Ecofeminist perspectives on or commoning care are present in initiatives like the Feminist Table in South Africa (Fakier & Cock, 2018), childcare commons in Barcelona (Zechner, 2021), Jineolojî pedagogy in Rojava (Piccardi & Barca, 2022), and in everyday practices of collective modes of (Earth-)care around the world (Gabbert & Lang, 2019).

20.5

Toward feminist ecological economics

Materialist ecofeminisms and related streams of thought at the intersection of feminisms and the environment (for an overview, see Dengler & Strunk, 2022) offer considerable insights for ecological economics and have fed into the development of feminist ecological economics (e.g., Perkins, 2007; O’Hara, 2009; Nelson & Power, 2018). In emphasizing the structural similarity of the externalization and exploitation of the “underwater part of the iceberg,” ecofeminisms around the world foreground intersecting power relations that, apart from destructive societal relations with nature, also include patriarchy, coloniality/

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racism, and class relations, thereby opening space for systemic alternatives. As (feminist) ecological economists, we should draw upon these insights and discuss, for example, the patriarchal and colonial roots of the growth paradigm. We should be aware that gender relations are omnipresent and that, more often than not, putatively gender-neutral policies reproduce intersectional inequalities. We should embrace the often-neglected feminist and post-/decolonial roots of critical environmental thinking and critically reflect on why this scholarship is often “structurally forgotten.” The feminist and the ecological critique of economics and growth have a lot in common, and feminist ecological economics, as a project in the making, can benefit from an increased theoretical integration of the two streams of thought. Corinna Dengler

Notes 1.

2.

3.

4.

This goes along Kelly Coate’s (2000: 177) claim that “feminism can no longer be spoken of in the singular, and any woman who does is immediately asked what she means and who she is speaking for.” It is noteworthy that also ecofeminist scholarship that leans toward this cultural strand (e.g., the work of Vandana Shiva) has more nuanced analyses than this critique suggests. For example, Shiva’s often criticized notion of a “feminine principle”/Prakriti is non-binary and dialectical rather than “biologically essentialising.” The term “intersectionality” has been coined by Kimberlé Crenshaw (1989) and the concept has firm (and much older) roots in Black feminist scholarship. For example, the Combahee River Collective (1977: X) defines what is later to be known as “intersectionality” by stating: “We are actively committed to struggling against racial, sexual, heterosexual, and class oppression, and see as our particular task the development of integrated analysis and practice based upon the fact that the major systems of oppression are interlocking.” The two icebergs resemble each other in distinguishing the visible, monetized “economy” from its invisible (or invisibilized), non-monetized, and putatively non-economic foundations. They differ, however, with regard to how the two spheres relate to each other – for a comparison of the two icebergs, see Collard and Dempsey (2020).

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Desire. Bloomington and Indianapolis: Indiana University Press. Nelson, Julie, Power, Marilyn (2018): Ecology, Sustainability, and Care: Developments in the Field. Feminist Economics 24(3), 80–88. O’Hara, Sabine (2009): Feminist Ecological Economics: Theory and Practice. In: Salleh, Ariel (ed.): Eco-Sufficiency and Global Justice: Women Write Political Ecology. London: Pluto Press, 180–96. Oksala, Johanna (2018): Feminism, Capitalism, and Ecology. Hypatia 33(2), 216–34. Patel, Raj, Moore, Jason (2017): The History of the World in Seven Cheap Things. Berkeley: California University Press. Pérez Orozco, Amaia (2014): Subversión Feminista de la Economía. Madrid: Traficantes de Sueños. Perkins, Patricia Ellie (2007): Feminist Ecological Economics and Sustainability. Journal of Bioeconomics 9(3), 227–44. Piccardi, Eleonora, Barca, Stefania (2022): Jin-Jiyan-Azadi: Matristic Culture and Democratic Confederalism in Rojava. Sustainability Science 17, 1273–85. https://​doi​ .org/​10​.1007/​s11625​-022​-01099​-x Plumwood, Val (1993): Feminism and the Mastery of Nature. London: Routledge. Quiroga Díaz, Nathalie (2015): Decolonial Feminist Economics: A Necessary View for Strengthening Social and Popular Economy. Viewpoint Magazine, October 31. Salleh, Ariel (1997): Ecofeminism as Politics: Nature, Marx and the Postmodern. London: Zed Books. Salleh, Ariel (2009) (ed.): Eco-Sufficiency & Global Justice: Women Write Political Ecology. London: Pluto Press. Schild, Verónica (2019): Feminisms, the Environment and Capitalism: On the Necessary Ecological Dimension of a Critical Latin American Feminism. Journal of International Women’s Studies 20(6), 23–43. Segato, Rita (2014): Colonialidad y patriarcado moderno: expansión del frente estatal, modernización, y la vida de las mujeres. In: Espinosa Miñoso, Yuderkis, Gómez Correal, Diana, Ocho Muñoz, Karina (eds.): Tejiendo de otro modo. Feminismo, epistemología y apuestes descoloniales en Abya Yala. Popayán: Editorial Universidad del Cauca, 75–91. Seymour, Nicole (2020): Queer Ecologies and Queer Environmentalism. In: Somerville, Siobhan (ed.): The Cambridge Companion to Queer Studies. Cambridge: Cambridge University Press, 108–21. Singh, Neera (2013): The Affective Labor of Growing Forests and the Becoming of Environmental Subjects: Rethinking

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Bauhardt, Christine, Harcourt, Wendy (eds.): Feminist Political Ecology and the Economics of Care. In Search of Economic Alternatives. London: Routledge, 55–69. von Werlhof, Claudia (1988): The Proletarian is Dead: Long Live the Housewife. In: Werlhof, Claudia, Mies, Maria, Bennholdt-Thomsen, Veronika (eds.): Women: The Last Colony. London: Zed Books, 168–81. von Winterfeld, Uta (2006): Naturpatriarchen: Geburt und Dilemma der Naturbeherrschung bei geistigen Vätern der Neuzeit. Berlin: oekom. Walsh, Catherine (2010): Development as Buen Vivir: Institutional Arrangements and (De) Colonial Entanglements. Development 53(1), 15–21. Warren, Karen (1996): Ecological Feminist Philosophies. Bloomington: Indiana University Press. Zechner, Manuela (2021): Commoning Care & Collective Power. Wien: Transversal.

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21. Ecological distribution conflicts This concept brings together ecological economics and political ecology. Ecological economics studies the metabolism of the industrial economy (i.e. the increasing and changing amounts of energy and materials entering the economy and exiting as waste). The procurement of energy and materials causes conflicts at the commodity extraction frontiers. Waste disposal also causes conflicts. The study of such conflicts is done by political ecology. These conflicts are also “valuation contests” (i.e. the protagonists of the conflicts display different valuation languages). Some would insist on the economic valuation of negative “externalities” while others appeal to ecological values, to livelihood needs, or to the sacredness of items of nature, such as special animals or trees, rivers, or mountains. These values are not commensurate. The increasing number of ecological distribution conflicts around the world is ultimately caused by the growing and changing metabolism of the economy in terms of growing flows of energy and materials. Changes in social metabolism and “ecological distribution conflicts” (EDC) are two sides of the same coin (Martínez-Alier, 1995; Martínez-Alier & O’Connor, 1996). EDC is a term for environmental injustices that originates from ecological economics. It has been used since 1995 to describe social conflicts born from the unfair access to natural resources and the unjust burdens of pollution. Environmental gains and losses are distributed in a way that causes such conflicts. EDCs give birth to movements of resistance, to the point that we can speak of a global movement for environmental justice. Many of such conflicts are gathered in the EJAtlas (http://​www​.ejatlas​.org), which in May 2023 is reaching 3880 cases registered. They must be seen as a large sample of a much larger quantity of EDC around the world and in recent history. The purpose of the EJAtlas is to research and help the world movements for environmental justice. The EDCs, if they obtain successful outcomes (stopping coal-fired power plants, nuclear power plants, hydropower dams, palm oil or eucalyptus

plantations), contribute to move the economy in a less unsustainable direction. For instance, a factory may be polluting the river (which belongs to nobody or belongs to a community that manages the river – as studied by Ostrom, 1990, and her school). This is not a damage valued in the market. The same happens with climate change, causing, perhaps already, sea level rise in some Pacific islands, or in the Kuna islands in Panama or the Sundarbans. Equally, a copper or bauxite smelting factory or a coal-fired power plant will pollute the air, perhaps causing respiratory illnesses. Such conflicts take place at local scales but also at national and international scales. More than market failures (a terminology that in mainstream economics implies that such externalities could be valued in money terms and internalized into the price system), these are “cost-shifting successes” (Kapp, 1950), which oftentimes lead to complaints from those bearing them. In these complaints, incommensurable values are deployed. For instance, if a “sacred grove” or small forest belonging to a tribal community in India is destroyed by open cast coal mining, financial compensation may be a way out for the company responsible, but other valuation languages (biodiversity, the “rights of nature”, human rights, the livelihood of local populations, Indigenous territorial rights, sacredness) will then be sacrificed. Sand mining in riverbeds is sometimes opposed by appealing to environmental values or local livelihoods or legal regulations, and sometimes by appealing also to the sacredness of rivers. Incommensurability of values means that they can be measured, but not in the same units. While the term “economic distribution conflicts” in political economy describes conflicts between capitalists and labour (profits versus salaries) or between landlords and peasants (over land rents), or conflicts on prices between sellers and buyers of commodities other than land or labour, or conflicts on the interest rate to be paid by debtors to creditors, the term EDC in political ecology stresses the idea that the unequal or unfair distribution of environmental goods and evils is not always coterminous with economic distribution. Without denying that economic compensation (economic valuation and payments for environmental services or for environmental

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liabilities) are appropriate sometimes, particularly in court cases after damage has been done already, the fact is that economic valuation and compensation means, in practice, the exclusion of social values that cannot be expressed in money terms, or that the protagonists of the conflicts prefer to express in other terms. For instance, impairment to health from pesticides, such as DBCP or chlordecone in banana plantations, can certainly be expressed in terms of genetic damage, or years of labour lost and of taking care of illness and disability in adults and old people. Such damage can be translated into money terms. However, some people might much prefer to count the damage also directly, without money translation, in units of human health and life, not least because we know that “the poor sell cheap”. In any case, who has the right, or rather the naked power, to impose one particular valu-

ation method when several incommensurable values are in dispute? Joan Martínez-Alier

References

Kapp, K. W. (1950). Social Costs of Private Enterprise. New York: Schocken Books. Martínez-Alier, J. (1995). The environment as a luxury good or “too poor to be green”?. Ecological Economics, Vol. 13(1): 1–10. Martinez-Alier, J., O’Connor, M. (1996). Ecological and economic distribution conflicts, in: R. Costanza, J., Martinez-Alier, O. Segura (Eds.), Getting down to Earth: Practical Applications of Ecological Economics. Washington DC: Island Press/ISEE, pp. 153–183. Ostrom, E. (1990). Governing the Commons. The Evolution of Institutions for Collective Action. Cambridge: Cambridge University Press.

Joan Martínez-Alier

22. Ecological macroeconomics Ecological macroeconomics is an approach to macroeconomics that explicitly recognizes the fundamental dependency of the economic activities of production, distribution, and consumption on materials and energy obtained from the planet of which the economy is a subsystem. These materials and energy may accumulate temporarily within the economy; they may, in the case of materials, be reused and recycled; but eventually they are returned to the environment as degraded waste, according to the physical principle of entropy. Associated with these processes is the extensive conversion of land, the destruction of the habitat for other species, and the consequent reduction in biodiversity. The rapid growth of materials and energy throughput and land conversion raises questions about the adequacy of sources of supply and the capacity of ecological systems to assimilate wastes, issues that are addressed in ecological macroeconomics. This account of ecological macroeconomics begins with its origins, including the development of some of its defining components, followed by an overview of recent research in ecological macroeconomics with an emphasis on models. It concludes with a set of research questions that give some idea of possible future directions for ecological macroeconomics.

Origins

The idea that economies are fundamentally dependent on nature has a long history. Francis Quesnay, a French economist and physician, published the Tableau Économique in 1759, nearly 20 years before Adam Smith’s celebrated Wealth of Nations. In his Tableau, Quesnay traced the income flows among the three sectors of the economy: landowners, agricultural workers, and artisans and merchants. He considered that only agriculture workers were truly productive, and the source of their productivity was the land on which they farmed. All other economic activities were seen as dependent on the surplus from agricultural production. At the time Quesnay was writing, France was essentially an agricultural economy

without capitalists. This was different in Britain, where the founders of classical economics – Smith, Malthus, and Ricardo – could already see the early stages of capitalism. Nonetheless, they continued to pay attention to the dependency of economies on land, and each of them thought that eventually economic expansion would cease. Malthus and Ricardo especially emphasized the limited capacity of land to provide sufficient food for a growing population. Later in the 19th century, with the continued rise of capitalism in Britain and some other countries, economists paid less and less attention to land, and to nature in general, such that in the 20th century it became quite common for economists to ignore nature altogether. They were content to assume that production depended only on labour, capital, and technology. Any limits arising from land or natural resources, it was assumed, could be overcome by increased capital and improved technology. In the 1960s and 1970s there was an upsurge of interest in the degradation of the environment, particularly in North America and some European countries. Environmental economics was born based on the principles of microeconomics and market failure, and on the concept of externalities introduced by Arthur Pigou in the 1920s. Markets can fail for many reasons. Externalities, which are effects on third parties not involved in market transactions, were seen as especially relevant to environmental degradation. Smoke from factories was an obvious example. Those who suffer its ill effects are not compensated financially in the normal course of doing business. To fix the problem, environmental economists proposed imposing a price on damaging emissions to air and water, and wastes. ‘Getting the prices right’ became the mantra. This could be done directly through a government-imposed tax on emissions, or indirectly through a system of emissions trading. The same logic was applied to resource problems such as overfishing, which is an example of the ‘tragedy of the commons’ discussed by Garrett Hardin (1968) that also applies to the overuse of air and water for disposing of wastes. With few exceptions (e.g. Victor 1972a, 1972b), most economists interested in environmental and resource problems in these early years conceived of them as problems of theoretical and applied microeconomics.

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Macroeconomics – the economics of the whole economy – was not considered relevant. Three who took a different view, but who did not use the term ecological macroeconomics, were Kenneth Boulding, Nicolas Georgescu-Roegen, and Herman Daly. In his short essay titled ‘The Economics of the Coming Spaceship Earth’, Boulding (1966) included many of the principles and concepts of ecological economics: economies as open and closed systems and their requirements for materials, energy, and knowledge, without which they cannot function; the thermodynamic principle of entropy and its very special implications for energy, which can be used more efficiently but cannot be recycled; and the principle that throughput of materials and energy should be minimized rather than maximized. Boulding also suggested that the obsession with production and consumption distorts technological change in a way that is counter to human welfare. Georgescu-Roegen (1971) provided an in-depth analysis of the implications of the entropy principle, also known as the second law of thermodynamics, for economics. His book The Entropy Law and the Economics Process remains the touchstone for all subsequent work in this area. Daly popularized Georgescu-Roegen’s ideas, which informed his own work on the steady-state economy. Daly (1968) also showed how a conventional input–output table could be extended to include the material and energy inputs and waste outputs that linked the economy to the natural environment. This was the birth of environmentally extended input–output analysis, first implemented empirically at the national level by Victor (1972/2018) and subsequently applied globally in multi-regional versions (e.g. Duchin 1998; Timmer et al. 2015). It has become a standard method for estimating the resource use and air, water, and solid waste associated with the purchase of goods and service produced in an economy, and has been incorporated in some ecological macroeconomic models. Daly (1989) also participated in the development of environmental accounting and alternatives to gross domestic product (GDP) to measure economic success, both of which became part of ecological macroeconomics.

Ecological macroeconomics – the first three decades

The foundations for ecological macroeconomics were laid by Herman Daly, who called for an ‘environmental macroeconomics’ in 1991. Although he did not use the specific term, Daly’s 1991 chapter is a good place to begin an account of what ecological macroeconomics is and what distinguishes it from other approaches to macroeconomics. Ecological macroeconomics stems from a pre-analytic vision of an economy different from other approaches. The pre-analytic vision of mainstream macroeconomics is the familiar circular flow of income graphic found in most textbooks. In its simplest form, the circular flow of income shows money flowing between households and firms in one direction, and goods and services, land, labour, and capital flowing in the other. This graphic is sometimes embellished by including government and banks and the associated flows of funds. What distinguishes the pre-analytic vision of ecological macroeconomics is the inclusion of the essential material and energy exchanges between the macroeconomy and the environment, which make economic activity possible. This is shown graphically by situating the circular flow of income within the planetary environment, and with material and energy flows between the planet and the macroeconomic subsystem. Energy flows between the planet and outer space: incoming solar energy and outgoing waste heat, complete the picture. Pre-analytic visions play a critical role in determining what is analysed. A vision of the macroeconomy as an isolated system, independent of material and energy ‘throughput’, excludes by default any consideration of constraints on economic growth from the depletion of natural resources and the disposal of wastes back to the environment. As Daly wrote in 1991 ‘the physical exchanges crossing the boundary between the total ecological system and the economic subsystem constitute the subject matter of [ecological] macroeconomics’ (35). The physical magnitude of these flows in relation to the environment determines the scale of the macroeconomy. Since the Earth is essentially a closed system (i.e. it mostly exchanges energy, not materials, with outer space), this raises questions about the physical scale of the economy measured by these flows, or alternatively by Peter A. Victor

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the stock of people and their artefacts made from these flows and maintained by them. Can the macroeconomy become too large to serve the best interests of humans and other species? Is there an optimal scale of the macroeconomy, and if so, how can it be determined? Is there a biocentric optimum, based on the well-being of all species, that is different from an anthropocentric optimum, based only on human well-being? Daly also asked, if consumption and production at the level of individual households and firms have an optimal scale, why does the aggregate of these activities not have an optimal scale as well? These are questions that arise from the pre-analytic vision of ecological macroeconomics but not from mainstream economics or most other heterodox approaches to economics. Yet it has become increasingly clear that the scale of the human economy has become so large that ecosystems at all levels, from local to global, are in decline. A macroeconomics that understands the significance of the scale of the macroeconomy would seem essential in the 21st century. By explicitly including material and energy flows, compliance of economics with the first and second laws of thermodynamics becomes critical in ecological macroeconomics. The first law relates to the quantities of materials and energy, which are conserved through all economic processes and the economy as a whole. The second law relates to the decline in the capacity of energy to do work whenever it is used. This is a decline in quality while quantity is maintained. Materials are also degraded with use, though there is debate about whether and how this is related to the second law of thermodynamics. In any case, the macroeconomy requires a continual throughput of material and energy and it is the increasing size of this throughput that is putting excessive pressure on planetary systems. Ecological macroeconomics provides ways of understanding this situation, its causes, consequences, and possible solutions. Ecological macroeconomics is not just about scale. It is also about distribution and allocation, about which much has been written by economists and others from widely different perspectives and which informs ecological macroeconomics. Daly follows Tinbergen’s (1952) principle of economic policy and calls for scale, distribution, and allocation each to have their own policy objective and instrument: scale should be susPeter A. Victor

tainable, distribution should be equitable, and allocation should be efficient. He also suggests a hierarchy such that scale is decided first (limits on throughput), followed by distribution (maximum and minimum incomes), and then allocation (through various mechanisms including markets). Daly ends his seminal paper on ecological macroeconomics with four principles for sustainable scale which have become part of a larger set of policy recommendations derived from ecological macroeconomics: 1) technical progress should be efficiency-increasing rather than throughput-increasing; 2) renewable resources should be harvested at rates not exceeding their rates of regeneration; 3) waste emissions should not exceed the renewable assimilative capacity of the environment; and 4) non-renewable resources should be exploited at a rate equal to the creation of renewable substitutes. The term ‘ecological macroeconomics’ first appeared in print in 2001, then again in 2008, in articles by Jonathan Harris (Daly had referred to it as ‘environmental macroeconomies’ while clearly taking a broader ecological perspective). In a paper titled ‘Ecological Macroeconomics: Consumption, Investment, and Climate Change’, Harris suggested that expenditures on consumption, investment, and by government, which constitute most of GDP, could each be divided into categories relating to their environmental impact (he also included a category of investment in human capital). He argued that economic growth could and should be reoriented to be consistent with ecological sustainability by concentrating growth in the categories of GDP with low energy intensity, in energy-conserving investment, and in government investment and human capital (Harris 2001, 2008, 2009). When examined empirically using UK data and the ecological footprint rather than energy as the environmental metric, Harris’s proposal was found to be valid, but only for a few years. The UK’s ecological footprint would decline for a few years with the change in the composition of GDP along the lines suggested by Harris, but after that, growth in GDP would overwhelm reductions in the footprint intensity (Victor 2012). The first, full-throated plea for an ecological macroeconomics using that name came from Tim Jackson. In his investigation into prosperity without growth (Jackson 2009), he devoted an entire chapter and appendix

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to the development of an ecological macroeconomics. His perspective, developed independently of Harris, shared some of the same features, such as a differentiation among types of investment: investments that enhance resource efficiency and lead to resource cost savings (e.g. energy efficiency, waste reduction, recycling); investments that substitute conventional technologies with clean or low-carbon technologies (e.g. renewables); and investments in ecosystem enhancement (e.g. climate adaptation, afforestation, wetland renewal). Jackson went further than Harris, especially in relation to ecological macroeconomic modelling. He wrote favourably about Victor’s (2008) LowGrow model of a non-growing economy based on Canadian data, which showed how economic growth could be gradually eliminated while greenhouse gas emissions, unemployment, and poverty are reduced so that economic and social stability are maintained. The model was highly aggregated and did not include a financial sector, but it was a start, and its use of systems dynamics was unusual in the world of macroeconomics. To make further progress, Jackson proposed five specific aims for ecological macroeconomic modelling: ● to test the stability of different macroeconomies under exogenously defined carbon emission and energy resource constraints; ● to explore the potential for macroeconomies with high investment to consumption ratios; ● to explore the potential for macroeconomies with high public sector expenditure and investments; ● to explore the stability of macroeconomies with low or no consumption growth; and ● to explore the stability of macroeconomies with low or no aggregate demand growth. In the 2010–2020 period there was a minor explosion of work on ecological macroeconomics, with over 35 papers published, and no doubt many more to come. In 2011, Røpke made an important contribution calling for ecological macroeconomics to have a broad scope, addressing core global and national challenges and problems: environmental, poverty and inequalities, large-scale migration, security implications of shifting global power relations, global economic crises

and instability of the financial system, and the complexity of global interrelationships. She also noted several challenges facing the construction of ecological macroeconomic models, few of which existed in 2011. Hardt and O’Neill (2017) surveyed the growing number of ecological macroeconomic models that had been developed to 2017. They identified three themes in the literature: 1) the need to manage an economy without growth, which is associated with the post-growth and de-growth literature, combined with improved understanding of what an economy is for; 2) a wider emphasis on developing new analytical methods and models that can represent the dependence of the macroeconomy on the natural environment, including how macroeconomic processes, such as unemployment, growth, and inflation, depend on natural resources and produce wastes, and how environmental damages feed back into the macroeconomy; and 3) the combination of post-Keynesian and ecological economics approaches with the need to build on insights from other heterodox economic fields (Hardt and O’Neill 2017). A useful distinction made by Hardt and O’Neill is between analytical and numerical ecological macroeconomic models. Analytical models contain few equations and can be solved analytically while numerical models contain many equations and use simulations to generate scenarios. Hardt and O’Neill divided the numerical models among four categories: stock-flow consistent, monetary input–output models, physical input–output models, and system dynamics. Many of the models included in their review fall into more than one of these categories. For example, Jackson and Victor’s GEMMA model (2015) is a systems dynamics simulation model that is stock-flow consistent with all monetary flows among sectors tracked and balance sheets maintained, and which also includes an input–output representation of production. The same is true of the more comprehensive ecological macroeconomic model of Dafermos et al. (2017). Their model is stock-flow consistent for monetary flows, physical stocks and flows satisfy the laws of thermodynamics, and stock-flow and fund-service resources are distinguished. Finance affects macroeconomic activity, and its amount and configuration determine Peter A. Victor

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material and energy throughput. Output is demand-determined, and allowance is made for climate change to affect its various components, while environmental damages and the exhaustion of natural resources can constrain supply. Jackson and Victor’s (2019, 2020) LowGrow SFC model is calibrated for a national economy and incorporates the equity and bond markets, the government sector, the central bank, and speculative housing investment, which are absent from the Dafermos model. Four categories of green investment are distinguished which have different macroeconomic effects depending on whether the investment is additional or displaces other investment, and whether it adds to the economy’s productive capacity or not. LowGrow SFC also includes a detailed electricity sub-model for simulating the conversion to renewable sources of electricity and the electrification of road and rail transportation but does not include an input–output component or the carbon cycle. Clearly, there are opportunities for enhancements to both models. Another difference between the Dafermos et al. and Jackson and Victor models is in the types of scenarios they have been used to explore. Dafermos et al. simulated the effects from 2015 to 2115 of assumptions about the sensitivity of economic activity to the leverage ratio of firms and different types of green finance policies on key environmental, macroeconomic, and financial variables. Jackson and Victor simulated four scenarios from 2017 to 2067: a base case, a carbon emissions reduction scenario, and two sustainable prosperity scenarios with non-growing GDP per capita, reduced work time, lower income inequality, declining environmental burden, stable household and government debt ratios, and stable unemployment. This difference in the choice of scenarios illustrates the different purposes to which such models are being put, not just for the exploration of post-growth possibilities but also for the analysis of the economic system as it exists today. Ecological macroeconomics is, by its nature, highly integrative. It combines in different ways the real economy, the financial economy, and natural systems. Fontana and Sawyer (2013, 2016) and Taylor et al. (2016) have explored the integration of ecological economics and post-Keynesian economics in considerable depth, highlighting their Peter A. Victor

similarities and differences. Rezai and Stagl (2016) have suggested that ecological macroeconomics should use the insights from other heterodox economic fields in addition to post-Keynesian approaches, such as Marxist, neo-Ricardian, and evolutionary economics. Røpke (2016) proposes the combination of three complementary system perspectives – the socio-technical provision systems, distributional systems, and macroeconomic systems – to address interrelated environmental, economic, and social crises. And Sers (2021) has demonstrated analytically and numerically how a dynamic input–output model, with endogenous money, prices, and energy return on energy invested (EROI) that is also stock-flow consistent and satisfies the first and second law of thermodynamics, can be combined with an integrated assessment model of climate change.

Future directions

Victor and Jackson (2020) suggest a research agenda for the further development of ecological economics under the headings: modelling, metrics, and contemporary issues. They propose ten specific research questions based on an assessment of the current state of ecological macroeconomics and where the greatest gains are to be made over the next decade: 1. How can the various ecological macroeconomic modelling approaches be improved and better integrated? 2. How can better use be made of a wider variety of databases covering all dimensions of relevance to ecological macroeconomics? 3. Building on the principles of stock-flow consistency, what improvements can be made to sub-models of the financial sector? 4. How can the representation of price formation and price-induced behaviours be improved within ecological macroeconomics? 5. What spatial and temporal scales are most appropriate for various types of ecological macroeconomic problems? 6. What lessons from outside economics should be incorporated in a world view that is more suitable as a framework for ecological macroeconomics?

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7. How can mediated modelling be used more effectively in the exploration and promotion of transitions to an economy that delivers sustainable prosperity for all? 8. What can be done to enhance the interactivity and accessibility of ecological macroeconomic models? 9. How can non-monetary metrics be improved and better integrated into ecological macroeconomics and models? 10. Are there growth imperatives in modern capitalism? Can they be overcome? If not, what are the broader implications for capitalism and the economic system itself? The extent to which these questions are addressed and answered will largely determine the course of ecological macroeconomics over the next decade. Peter A. Victor

References

Boulding K. E. (1966), ‘The Economics of the Coming Spaceship Earth’. In Jarrett, H. (Ed.), Environmental Quality in a Growing Economy: Essays from the Sixth RFF Forum, 3–14. Johns Hopkins University Press. Dafermos, Y., Nikolaidi, M., and Galanis, G. (2017), ‘A Stock-Flow-Fund Ecological Macroeconomic Model’, Ecological Economics, 131, 191–207. Daly, H. E. (1968), ‘Economics as a Life Science’, Journal of Political Economy, 76, 392–406. Daly, H. E. (1989), ‘Toward a Measure of Sustainable Social Net National Product’, in Y. J. Ahmad, S. El Serafy, and M. Lutz (Eds.), Environmental Accounting for Sustainable Development, 8–9. World Bank. Daly, H. E. (1991), ‘Elements of Environmental Macroeconomics’, in R. Costanza (Ed.), Ecological Economics: The Science of Management and Sustainability and Management, 32–46. Columbia University Press. Duchin, F. (1998), Structural Economics: Measuring Change in Technology, Lifestyles and the Environment. Island Press. Fontana, G., and Sawyer, M. (2013), ‘Post-Keynesian and Kaleckian Thoughts on Ecological Macroeconomics’, European Journal of Economics and Economic Policies: Intervention, 10, 256–67. Fontana, G., and Sawyer, M. (2016), ‘Towards Post-Keynesian Ecological Macroeconomics’, Ecological Economics, 121, 186–95. Georgescu-Roegen, N. (1971), The Entropy Law and The Economic Process. Harvard University Press.

Hardin, Garrett. (1968), ‘The Tragedy of the Commons’, Science, 162(3859), 1243–8. Hardt, L., and O’Neill, D. W. (2017), ‘Ecological Macroeconomic Models: Assessing Current Developments’, Ecological Economics, 134, 198–211. Harris, J. (2001), ‘Macroeconomic Policy and Sustainability’, Global Development and Environment Institute Working Paper No. 01-09. Harris, J. (2008), ‘Ecological Macroeconomics: Consumption, Investment, and Climate Change’, Global Development and Environment Institute Working Paper No. 08-02. Harris, J. (2009), ‘Ecological Macroeconomics: Consumption, Investment, and Climate Change’, Real World Economics Review, 50, 34–48. Jackson, T. (2009), Prosperity without Growth. Earthscan. Jackson, T., and Victor, P. A. (2015), ‘Toward an Ecological Macroeconomics’, in P. Brown and P. Timmerman (Eds.), Ecological Economics for the Anthropocene, 237–59. Columbia University Press. Jackson, T., and Victor, P. A. (2019), ‘Managing Without Growth: Exploring Possibilities’, in P.A. Victor (Ed.), Managing Without Growth. Slower by Design, Not Disaster, 2nd ed., 271–303. Edward Elgar Publishing. Jackson, T., and Victor, P. A. (2020), ‘The Transition to a Sustainable Prosperity: A Stock-Flow-Consistent Ecological Macroeconomic Model for Canada’, Ecological Economics, 177, 106787. Quesnay, F. (1972 [1759]), Tableau économique, 3rd ed. M. Kuczynski and R. Meek, Eds. Macmillan. Rezai, A., and Stagl, S. (2016), ‘Ecological Macroeconomics: Introduction and Review’, Ecological Economics, 121, 181–5. Røpke, I. (2011), ‘Ecological Macroeconomics – Calling for a Shift from Consumption to Investment’, Socio-Technical Transitions, Social Practices, and the new Economics: Meeting the Challenges of a Constrained World. Paper for the SCORAI Workshop, “Socio-Technical Transitions, Social Practices, and the New Economics: Meeting the Challenges of a Constrained World”, Princeton, NJ, 15–16 April. Røpke, I. (2016), ‘Complementary System Perspectives in Ecological Macroeconomics — The Example of Transition Investments During the Crisis’, Ecological Economics, 121, 237–45. Sers, M. (2021), ‘Towards an Ecological Macroeconomics: Linking Energy and Climate in a Stock-Flow Consistent Input– Output Framework’, PhD Dissertation, York University.

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132  Elgar encyclopedia of ecological economics Taylor, L., Rezai, A., and Foley, D. (2016), ‘An Integrated Approach to Climate Change, Income Distribution, Employment, and Economic Growth’, Ecological Economics, 121, 196–205. Timmer, M. P., Dietzenbacher, E., Los, B., Stehrer, R., and de Vries, G. J. (2015), ‘An Illustrated User Guide to the World Input–Output Database: The Case of Global Automotive Production’, Review of International Economics, 23, 575–605. Tinbergen, J. (1952), On the Theory of Economic Policy. North-Holland. Victor, P. A. (1972a), ‘The Macro-Economics of Pollution,’ Technology and Society, 7(4), 118–21.

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Victor, P. A. (1972b), Economics of Pollution. Macmillan. Victor, P. A. (2018/1972), Pollution: Economy and Environment. Routledge. Victor, P. A. (2008), Managing Without Growth. Slower by Design, not Disaster. Edward Elgar. Victor, P. A. (2012), ‘Growth, Degrowth and Climate Change: A Scenario Analysis’, Ecological Economics, 84, 206–12. Victor, P. A., and Jackson, T. (2020), ‘A Research Agenda for Ecological Macroeconomics’, in R. Costanza, J.D. Erickson, J. Farley, I. Kubiszewski (Eds.), Sustainable Wellbeing Futures: A Research and Action Agenda for Ecological Economics, 357–72. Edward Elgar.

23. Ecological unequal exchange Introduction

Ecological Unequal Exchange (EUE) is a direct heir to unequal economic exchange. The latter arises as a structuralist critique of David Ricardo’s (1817/1950) theory of comparative advantage that has fostered the belief that international trade (IT) is always advantageous for the countries that participate in the exchange. In this perspective, trade is conceived as a positive-sum game where all participants are winners (Pérez-Rincón, 2006). For traditional economists, all types of exchange are fair since they are defined in the exclusive sphere of the market by matching the price that the buyer is willing to pay and the seller is willing to accept (Howell, 2007). In contrast, the structuralist approach sees free trade, embedded in an unequal institutional, cultural, and power relations structure, as a negative-sum game where there are winners and losers in monetary terms (Shaikh, 1980/2008; Andersson and Lindroth, 2001). The winners correspond to the “Northern” countries exporting capitaland knowledge-intensive goods. The losers are the countries of the “South”,1 specialised in exporting raw materials and unskilled labour-intensive goods. Both approaches extend their logic to the relationship between IT and the environment. The orthodox view, based on the Environmental Kuznets Curve, considers that free trade also promotes environmental sustainability. Economic growth resulting from free trade improves both the amount of economic resources available for environmental protection activities and society’s acceptance of increased spending on these activities (Van Hauwermeiren, 1998). For its part, by incorporating biophysical flows into the analysis, Ecological Economics (EE) questions these advantages in two directions: the scale effect and the equity effect. In the first, trade liberalisation is an important factor in the dynamics of global environmental deterioration due to the increase in the amount of material and energy resources that are mobilised with the growth of trade in a world of finite resources (International Resource Panel, 2019). In the second, trade is not a positive-sum game in

environmental terms between trading countries given the material imbalance produced by the exchange between countries importing material and energy resources (industrialised) and countries exporting these types of goods and importing manufactures and knowledge, the countries of the South (Martínez-Alier, 2007).

Conceptual foundations of EUE

One of the ontological foundations of EE is “critical realism”. This postulates “the existence of an objective reality independent of human beings”. Thus, unlike other social sciences, EE is primarily concerned with biophysical reality (Spash, 2012). The biophysical perspective extends to the analysis of IT, showing that its asymmetries not only occur at the level of monetary flows, but are also reflected in biophysical flows. In this sense, exchange is not only economically unequal, but also ecologically unequal. An EUE cannot be captured by conventional economic studies that analyse only monetary phenomena. It requires considering the biophysical aspects of production, transport, and consumption, where the Second Law of Thermodynamics becomes essential (Hornborg, 2012; Martínez-Alier, 2007). When comparing price and mass evolution in the value chain of a product, the Second Law of Thermodynamics implies that as one moves up the production chain, entropy irreversibly increases and, at the same time, the exergy (energy available for work) of the original inputs decreases. The loss of initial mass and productive potential is accompanied by an increase in prices as one moves up the production chain due to the higher economic value-added incorporated (Dorninger and Eisenmenger, 2016). The inverse relationship in which raw materials and energy have a low economic value while processed goods that have already dissipated much of their energy and materials have a high monetary value (Hornborg, 1998a; Pérez-Rincón, 2006) explains the IT between the Centre and the Periphery. Its objective is none other, in metabolic terms, than the transfer of energy and other resources from the peripheries to the centres of accumulation. The price differential allows for international exchange, but this exchange is ecologically unequal as it transfers entropy from the countries of the Centre to the Periphery (Hornborg, 2012).

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However, EUE cannot be explained by biophysical realities alone as a natural process. Economic and ecological inequities are constructed and driven by politico-economic actors, including countries, institutions, and companies with great power on the international stage. Thus, it is asymmetrical power relations that promote the development of these monetary and biophysical imbalances through norms and institutions that shape and structure unequal exchange processes. The political perspective associated with power asymmetries is recognised by all scholars of EUE, not only metabolic, but also economic. For example, Muradian and Martínez-Alier (2001, 294, cited by Rice, 2007, 53) point out that: “undervaluation has less to do with market failures than with the successful appropriation of natural resources by the most powerful trading partners, without the internalisation of all ecological and social costs”. Jorgenson (2016) states that although EUE is, by definition, the analysis and implications of the asymmetric flow of biophysical (matter-energy) resources between the periphery and the core of the world system, behind these unequal flows are systems of unequal political and power relations. An important achievement of the literature on EUE is that it has succeeded in exposing that prices reflect power structures, and that the market value of biophysical volumes of resources varies according to these structures (Howell, 2007). This dynamic is further aided by aggressive international competition driven by globalisation, which has fuelled a so-called “race to the bottom” that has contributed to the dismantling of existing regulatory norms by pushing down social and environmental standards (Rudra, 2008). At this point it is necessary to incorporate the concept of “environmental costs promoted or transferred” to other sites, extending the scale of the analysis beyond national borders (Muradian et al., 2002). From this perspective, IT can be seen as a new “environmental vector” (like air and water) that spreads ecological burdens and impacts between countries across borders (Karlson, 1995). The increased physical and social “distance” associated with trade, between decision-makers and those who suffer from them, makes it difficult for people to see the consequences of their actions (Daly, 1993).

Mario Pérez-Rincón

History of EUE

The background of EUE can be traced back to the history of unequal economic exchange. This idea has a long tradition and has its origins in the contributions of the classical economists of the 19th century and Karl Marx. The Marxist theory of unequal exchange has its beginnings in Otto Bauer (1881–1938), who examined trade relations in the context of the Austro-Hungarian Empire. He found non-equivalent values in the exchanged commodities that favoured regions with a higher organic composition of capital (Bauer, 1924/2000, 200). Subsequently, Emmanuel identified the origin of unequal exchange in a transfer of value from countries with lower wage and organic capital compositions, produced by the contrast between capital mobility versus almost no labour mobility preserving the wage disparity (Emmanuel, 1969). This concept, understood as exchanging more embodied labour for less, became, within Marxist thought, an important factor in the anti-imperialist and decolonisation struggle of most of the Global South. This perspective was strengthened by world systems analysis: the trade-induced global division of labour gives rise to an international structure of unequally powerful nation states. Such a structure determines an accelerated process of accumulation in certain regions (the core), while imposing a cycle of backwardness on others (the periphery; Wallerstein, 1974; Frank, 1966/2008). However, it is Raúl Prebisch (1949) together with Singer (1950), who is credited with founding the theory of economically unequal exchange in IT (Oulu, 2015). It is therefore referred to as the Prebisch– Singer thesis. For Prebisch, leader of the Latin American dependency theory and first director of the Economic Commission for Latin America and the Caribbean (ECLAC), the maldistribution of the benefits of IT is attributed to differences both in the income elasticities of demand for primary products and manufactured goods, and to asymmetries in the functioning of labour markets that limited the effectiveness of trade union action in wage determination (Prebisch, 1949). But in addition, the unequal exchange moves over time against the countries producing primary goods. More and more primary products are required to buy the same amount of final goods from industry (Prebisch, 1949, 3). Like

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Marxist thought, Prebisch considered that external imbalances are directly associated with the hierarchical structure of the world economic system, where each group of countries plays a role in feeding that system, and where the uneven spread of technical progress among the countries of the world is facilitated by the deterioration of the terms of trade (Prebisch, 1952). Warnelius (2016) argues that, for its part, the development of an ecological theory of unequal exchange, defined as the exchange of more “natural values” for less, began with Stephen Bunker (1984, 1985) working in Brazil. He argued that: (i) differences in the economies of peripheral and core countries create unequal exchange in terms of labour embodied in products (economic), but in addition there is transfer and appropriation of energy and matter from the periphery to the centre (ecological); and (ii) the extraction and export of natural resources affect the further development potential of the periphery (Bunker, 1985). Subsequently, in 1988, Howard T. Odum raised the most famous theory of unequal exchange based on thermodynamics (Odum, 1988; Odum and Arding, 1991). This ecologist conceptualised EUE in terms of energy and argued that the North imports “embodied energy” (emergy), or “energy memory”, from the South, to explain the unequal exchange of energy between nations (cited by Howell, 2007). The inequality of this exchange arises from the periphery’s export of the energy content of its natural resources, which is not accounted for in prices: Odum believed that “the periphery is being underpaid for the energy memory content of its natural resources because they are free gifts of nature and therefore not properly valued in the market” (Hornborg 1998a, 131). During the 1990s, the work of Alf Hornborg (1998a) appeared; he conceptually improved the EUE theory by advocating the evaluation of net flows of matter-energy (productive potential), but without equating it with economic “value”. Hornborg also considers that it is better to work with the concept of “exergy”, which is the quality of energy or the part of it that is available for mechanical work. The finished products, in turn, contain a diminished productive potential but an increased economic value and utility, the control of which facilitates the acquisition of

even greater unprocessed raw materials and energy needed to maintain and expand the industrial technomass (Hornborg, 1998b). Andersson and Lindroth (2001) extend the EUE perspective, which focused more on resource and energy appropriation, to the pollution or environmental burden shifting from Northern to Southern countries resulting from IT. Southern resources such as raw materials or biodiversity can be used to absorb pollution (Andersson and Lindroth, 2001, 116). This broader view of EUE is also incorporated by Muradian et al. (2002) and by Stefan Giljum and Nina Eisenmenger (2003). By including both aspects (resources and pollution), Rice (2007, 44) broadens the concept: EUE can be defined as referring to the “increasingly disproportionate use of ecological systems and the externalisation of negative environmental costs by the core industrialised countries and, consequently, the diminishing opportunities for utilisation and the imposition of exogenous environmental burdens on the periphery”. This can also be seen as industrialised countries “importing” sustainability from poor countries and thus preserving their local ecological capital even if they consume more biomass and sink capacity than is produced within their own borders (Anderson and Lindroth, 2001). The displacement of extractive frontiers “elsewhere” is linked to the increase in environmental conflicts and the development of the global environmental justice movement (Martínez-Alier, 2020). In this sense, there is a clear relationship between IT, productive specialisation towards extractive sectors, and environmental conflicts, which accentuates their consequences in the territories of the countries of the South. The Environmental Justice Atlas (2022), the largest database on ecological conflicts in the world (https://​ ejatlas​ .org/​ ), demonstrates this reality for Latin America and the Caribbean (LAC) by showing that most of the conflicts are generated in the extractive export sectors (minerals, biomass, and fossil energy).

Methods to identify the EUE

The metabolic aspects involved in the analysis of EUE make it necessary to use biophysical methods and indicators of sustainability to identify it. The main tool used to understand the inequality of trade is the material and energy balances of trade between countries Mario Pérez-Rincón

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or groups of countries. Without biophysical balances it is impossible to identify EUE. The main biophysical indicators used to make these balances include: i) ecological, carbon, and water footprints; and ii) material flow analysis (MFA), which is the most widely used method. This has been developed by the so-called Vienna School led by Marina Fischer-Kowalski (Fischer-Kowalski and Haberl, 1997). MFA provides a specific indicator to measure EUE: the biophysical trade balance, which is measured in tonnes (Dittrich and Bringezu, 2010; Fischer-Kowalski et al., 2011). Other indicators include: iii) energy balances; iv) human appropriation of net primary energy (Haberl, 1997); v) physical input–output analysis (PIOT; Moran et al., 2009); vi) life cycle analysis (LCA; Oulu, 2015); and vii) commodity chain analysis in the IT (Hughes, 2001). These last two methodologies allow identifying the EUE of a specific product.

Empirical expressions of EUE

The environmental inequality of the IT is produced through two effects. One is a net asymmetry in the material and energy balance against the countries of the South. This results in a permanent biophysical deficit that shifts most of the environmental pressures and burdens onto their territories, leading to the deterioration of their natural heritage. This EUE perspective is the most widely used and has ample empirical evidence at the level of nation states and groups of countries (Pérez-Rincón, 2006; Infante-Amate and Krausmann, 2019; Dorninger and Eisenmenger, 2016; Rivera-Basques et al., 2021; I ̇pek Tunç et al., 2022; Zhang et al., 2023). At the global level, a recent study was conducted for 163 countries using a multi-regional extended input–output model (Dorninger et al., 2021). This work tested the hypothesis that EUE is a persistent feature of the global economy between 1990 and 2015 for four groups of resources: materials, energy, land, and labour. For the case of LAC, there is evidence that the economies of the region are biophysically more open to outflows of material resources than to inflows (PNUMA-SCIRO, 2013). The longest-term study that has been done for the region (1900–2016), and perhaps globally, found

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that, “without exception, the region has been a net supplier of materials to the rest of the world; that is, its exports have always been greater than its imports in biophysical terms. Moreover, this biophysical deficit is growing, from 4 million metric tonnes (Mt) in 1900 to 610 Mt in 2016” (Infante-Amate et al., 2022, 9). The second effect is unfavourable terms of trade (UTT; ratio of export prices to import prices) against commodity-exporting countries, as Prebisch (1949) also pointed out for monetary exchange. In other words, products are exported without including in the prices the environmental damage generated, in addition to the damage to human health, which, in many cases, nature has taken a long time to produce. This can be called “ecological dumping” by selling below production costs (Martínez-Alier and Roca-Jusmet, 2018, 522). This perspective is used by dividing the tones exported or imported of a good or a group of goods by their monetary value. The UTT are easy to obtain as information on tones and monetary income or outflows from trade are reported annually in national and international statistics on IT. This path is rarely used in EUE analysis. However, there are some works that address the UTT perspective for some LAC countries, for example, Colombia (Pérez-Rincón, 2006), and for South America as a whole (Alonso-Fernández and Regueiro-Ferreira, 2022). In synthesis, EUE is a concept constructed by the EE to highlight the environmental inequities of the IT that affect Southern countries and benefit Northern countries. These environmental inequities are driven by: i) the asymmetry between the physical value of natural resources (rich in available energy) and their economic valuation (little added monetary value); and ii) the unequal power relations that allow, promote, and facilitate this type of exchange. Historically, EUE is heir to unequal economic exchange and has a similar horizon of denouncing, making visible and rejecting the existing power structures that generate inequalities. EUE can only be expressed in biophysical terms through material and energy balances, where its central expression is that more environmental values are exchanged for less. Mario Pérez-Rincón

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Note

1. The political economy analytical categories of “The North” and “The South” are used to capture the dynamics of EUE. The North refers to the economically and politically powerful countries geographically concentrated in the northern hemisphere. The South refers to the economically and politically disadvantaged countries of the world concentrated in the southern hemisphere (Howell, 2007). These categories also correspond to centre–periphery.

References

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their sustainability problems. Society & Natural Resources. 10 (1): 61–85. Fischer-Kowalski, M., Krausmann, M., Giljum, S., Lutter, S., Mayer, A., Bringezu, S., Moriguchi, Y., et al. (2011). Methodology and indicators of economy-wide material flow accounting. Journal of Industrial Ecology 15(6): 855–76. Frank, A. G. (2008 [1966]). The development of underdevelopment. In Seligson, M.A., and Passe-Smith, J. T. (eds.), Development and Underdevelopment: The Political Economy of Global Inequality. London: Lynne, Rienner Publishers, 257–68. Giljum, S., Eisenmenger, N. (2003, June). North– South trade and the distribution of environmental goods and burdens: A biophysical perspective. SERI Working Paper, No. 2. Haberl, H. (1997). Human appropriation of net primary production as an environmental indicator: implications for sustainable development. Ambio 26(3): 143–6. Hauwermeiren, S. V. (1998). Manual de Economía Ecológica. Programa de Economía Ecológica, Instituto de Ecología Política, Santiago de Chile. Howell, G. (2007). The North–South environmental crisis: an unequal ecological exchange analysis. New School Economic Review 2(1): 77–99. Hornborg, A. (1998a). Towards an ecological theory of unequal exchange: articulating world system theory and ecological economics. Ecological Economics 25: 127–36. Hornborg, A. (1998b). Ecosystems and world systems: accumulation as an ecological process. Journal of World-Systems Research 4: 169–77. Hornborg, A. (2012). Global Ecology and Unequal Exchange: Fetishism in a Zero-Sum World. London: Routledge. Hughes, A. (2001). Global commodity networks, ethical trade and governmentality: organizing business responsibility in the Kenyan cut flower industry. Trans Institutional British Geographers 26: 390–406. International Resource Panel (2019). Global Resources Outlook 2019: Natural Resources for the Future We Want. A Report of the International Resource Panel. Nairobi: United Nations Environment Programme. Infante-Amate, J., Krausmann, F. (2019). Trade, ecologically unequal exchange and colonial legacy: the case of France and its former colonies (1962–2015). Ecological Economics 156: 98–109. Infante-Amate, J., Urrego-Mesa, A., Piñero, P., Tello, E. (2022). The open veins of Latin America: long-term physical trade flows (1900–2016). Global Environmental Change 76: 102579. I ̇pek Tunç, G., Akbostancı, E. Türüt-Asık, S. (2022). Ecological unequal exchange between

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138  Elgar encyclopedia of ecological economics Turkey and the European Union: an assessment from value added perspective. Ecological Economics 192: 107269. Jorgenson, A. (2016). The sociology of ecologically unequal exchange, foreign investment dependence and environmental load displacement: summary of the literature and implications for sustainability. Journal of Political Ecology 32: 334–49. Karlson, R. (1995). Recycling in Life Cycle Assessments. Doctoral thesis, Chalmers University of Technology, Gothenburg. Martínez-Alier, J. (2007). Marxism, social metabolism, and international trade. In Hornborg, A., McNeil, J. R., and Martínez-Alier, J. (eds.), Rethinking Environmental History: World-System History and Global Environmental Change. Lanham, MD: AltaMira Press, 221–38. Martínez-Alier, J., Roca-Jusmet, J. (2018). Economía Ecológica y Política Ambiental, 3rd ed. México: FCE. Martínez-Alier, J. (2020). A global environmental justice movement: mapping ecological distribution conflicts. Disjuntiva 1(2): 81–126. https://​ doi​.org/​10​.14198/​DISJUNTIVA2020​.1​.2​.6 Moran, D. D., Wackernagel, M. C., Kitzes, J. A., Heumann, B. W., Phan, D., Goldfinger, S. H. (2009). Trading spaces: calculating embodied ecological footprints in international trade using a product land use matrix (PLUM). Ecological Economics 68: 1938–51. Muradian, R., Martínez-Alier, J. (2001). Trade and the environment from a “Southern” perspective. Ecological Economics 36: 281–97. Muradian, R., O’Connor, M., Martínez-Alier, J. (2002). Embodied pollution in trade: estimating the “environmental load displacement” of industrialized countries. Ecological Economics 41: 51–67. Odum, H. T. (1988). Self-organization, transformity, and information. Science 242: 1132–9. Odum, H. T., Arding, J. E. (1991). Emergy Analysis of Shrimp Mariculture in Ecuador. Coastal Resources Center, University of Rhode Island. Oulu, M. (2015). The unequal exchange of Dutch cheese and Kenyan roses: introducing and testing an LCA-based methodology for estimating ecologically unequal exchange. Ecological Economics 119: 372–83. Pérez-Rincón, M. (2006). Colombian international trade from a physical perspective: towards an ecological “Prebisch thesis”. Ecological Economics 59(4): 519–29. Prebisch, R. (1949). El desarrollo económico de la América Latina y algunos de sus princi-

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pales problemas. Santiago de Chile: ECLAC, E/CN.12/89. Published in English as The Economic Development of Latin America and Its Principal Problems (1950). New York: UNCLA. Prebisch, R. (1952). Theoretical and Practical Problems of Economic Growth (E/CN.12/221): Santiago de Chile, ECLAC. PNUMA-SCIRO (2013). Tendencias del flujo de materiales y productividad de recursos en América Latina. Número de trabajo: DEW/1578/PA, Panamá. Ricardo, D. (1950 [1817]). The Works and Correspondence of David Ricardo (Edition prepared by Piero Sraffa), Vol. 1: On the Principles of Political Economy and Taxation: London, Cambridge University Press. Rice, J. (2007). Ecological unequal exchange: consumption, equity, and unsustainable structural relationships within the global economy. International Journal of Comparative Sociology 48(1): 43–72. Rivera-Basques, L., Duarte, R., Sánchez-Chóli, J. (2021). Unequal ecological exchange in the era of global value chains: the case of Latin America. Ecological Economics 180: 106881. Rudra, N. (2008). Globalization and the Race to the Bottom in Developing Countries: Who Really Gets Hurt? Cambridge: Cambridge University Press. Shaikh, A. (2008 [1980]). On the laws of international exchange, in Nell, E. J. (ed.), Growth, Profits and Property. Cambridge: Cambridge University Press, 204–35. Singer, H. (1950): U.S. foreign investment in underdeveloped areas, the distribution of gains between investing and borrowing countries. American Economic Review 40: 473–85. Spash, C. (2012). New foundations for ecological economics. Ecological Economics 77: 36–47. Wallerstein, I. (1974). The Modern World System: Capitalist Agriculture and the Origins of the European World Economy in the Sixteenth Century. New York: Academic Press. Warlenius, R. (2016). Linking ecological debt and ecologically unequal exchange: stocks, flows, and unequal sink appropriation. Journal of Political Ecology 23: 364–80. Zhang, Y., Liao, C., Pan, B. (2023) Ecological unequal exchange between China and European Union: An investigation from global value chains and carbon emissions viewpoint, Atmospheric Pollution Research, 14(2): 101661. https://​doi​.org/​10​.1016/​j​.apr​.2023​ .101661.

24. Economic anthropology Economic anthropology (EA) is a subdiscipline to anthropology that displays considerable overlap of subject matter with ecological economics (EE). EA was initially influenced by turn-of-the-century economic and evolutionary theory, particularly that of K. Marx and F. Engels, inasmuch as they described pre-industrial economic and living conditions in England, India, Indonesia, and elsewhere. The Marxist focus on labour, productive forces, and means of production were conducive to describing the economics of the type of communities anthropologists are usually interested in (i.e. non-industrial, tribal, feudalistic, or egalitarian societies; Eriksen 2001). While this provided an early set of tools to describe non-industrial economies, deeper analysis became possible with influence of substantivist and institutional economist K. Polanyi. In his analyses of markets and economic transition, Polanyi was able to show that (a) markets are human-made, not guided by a make-believe invisible hand, and that (b) economic arrangements are based on cultural values, trust, and social arrangements that are negotiated among various segments of the population (Polanyi 1944, 1957). In particular, Polanyi’s analysis of markets uses a wealth of ethnographic data from various world regions. Additional critical influences to the evolution of EA were agricultural economist Ester Boserup (1965/2002), who linked population size and growth with economic transition, as well as sociologist Marcel Mauss (1950/2002), who highlighted the importance of reciprocity and redistribution of resources in traditional societies: institutional obligations and cultural morals create bonds that ultimately lead to the formation of communities; cultural symbols, such as money, express these bonds and communicate them within the community. Money, thus, essentially simplifies and standardises communication around relationships between actors in a community (Egbert 2018). Exposed to these influential writers, as well as the forefathers of modern anthropology (B. Malinowski, F. Boas, R. Firth), the onus was on two distinguished anthropologists, M. Harris (1979) and M. Sahlins (1974),

to further define the subfield of EA. While Harris followed the Marxian tradition in what he defined as cultural materialism, Sahlins followed the institutional route ploughed by Polanyi. Several basic assumptions were included in both schools of thought: firstly, that human economies depend on nature as much as they depend on social organisation; secondly, that production – or resource use – needs to be organised socially (i.e. there are various ways that this can be achieved); and thirdly, that the social conditions of production have an effect on the extent and intensity of the resource use (Seiser 2017). While there may be varying rationalities to explain which resources are used, how they are extracted, and into what they are transformed, early economic anthropologists were in unison about the fact that achieving maximum productivity was rarely the goal among pre- or non-industrial societies. For one, resource use decisions are rarely taken at the individual level. Rather, the focus has been on group decisions carried out by economic units, such as households, communities, tribes, clans, and so on. Geertz (1983) describes how decisions occur within a specific cultural setting and knowledge environment he terms common sense. Grünbühel and Williams (2016) build on this and identify livelihood strategies and risk evaluation as primary, socially constructed determinants for decisions. In addition, social institutions mediate communal decisions, which tends to normalise outliers through their inherent power in the community. In non-differentiated societies, wealth is often connotated with negative values. Mechanisms of reciprocity and redistribution (Mauss 1950/2002; Rapaport 1984/2000) have the function to ensure that both wealth and poverty do not go to extremes. For example, poor people in Thailand are welcome to stay at temples, should they require housing, and eat if they are hungry, while rich community members are expected to give more generously to the monks who then redistribute further to those in need. Furthermore, property is a term that varies in its meaning across cultures and many ethnographic examples show that access to resources does not automatically imply the possibility of amassing individual wealth. Similarly, common property resources are widespread, which, similar to Ostrom’s analogous insights (Ostrom &

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Hess 2007), regulate resource use and often prevent exploitation for the purpose of wealth amassment.1 Rather than maximising productivity and optimising transformation of goods, other values, such as spending time on social activities, may be in the forefront. This is, in part, because non-differentiated societies (mechanistic societies, according to the sociologist Durkheim, 1933/1997) possess weak formal regulatory institutions at the state level. These are compensated by strong social institutions at the community level. For example, the judiciary is often managed within the community without involving professional legal personnel. This is certainly the case among communities that practice kanun law in the Southern Balkans – a largely oral tradition mediated by community members that regulates, among other things, inheritance, resource control, and blood feud (Young 2001). Hence, non-differentiated societies spend more time in activities that regulate and reproduce the community’s institutions as well as its cultural and symbolic systems (Ellen 1990). Likewise, due to varying population structures, dependency ratios can be higher compared to industrial societies. Together with missing institutions of education, care, and social security, this requires the community to invest time into schooling the young and looking after the old and sick (Godelier 1975). The requirement to maintain community-level institutions leads to group interests dominating individual interests in these societies. Belief systems often support and regulate these group values, a fact first described by sociologist M. Weber (Eriksen 2001). More recent fields of investigation in EA include a view on unequal exchange among world regions (Silverman 2005); increasing dependency of localised economies on larger-scale influences, such as the state, world markets, and the importance of regional infrastructure on livelihood options (Scoones 2009); how economic hegemony is constructed using morals, values, symbols, and ideologies (based on Foucault’s (1970) work on how discourse is formed); as well as applying livelihood approaches to the study of local economies (see also, below).2 As the primary touch points to ecological economics and the two main schools of thought in EA, but also by merit of the fact that ecological economics has much to learn from studying Clemens M. Grünbühel

modern-day non-industrial economies, and due to the potential of EA to find renewed relevance by aligning closer to ecological economics, the remaining paragraphs look into major contributions of EA, followed by more recent developments in the field and a call for collaboration. Building on early ecological economics as well as cultural ecology (the study of human– nature interactions within the field of anthropology), cultural materialism establishes the fact that resource extraction, resource transformation, and resource exchange build the foundation of every economy, regardless of its scale and size. In fact, culture itself is built on this foundation. To manage and move resources, however, a degree of social organisation is necessary, which includes formal and informal institutions, management of relationships within a community, and the distribution of power (political economy) within the social system. Stacked on top of these, is the “superstructure” (i.e. ideological, religious, and symbolic ideas), which serves to set priorities, preferences, and decisions and subdues people into accepting the order of things. In the words of M. Harris, the three stacked levels of (a) resource utilisation, (b) social organisation, and (c) ideology are termed infrastructure, structure, and superstructure (Buzney & Marcoux n.d.; Harris 1979). Following the ideas of cultural ecology, cultural materialism maintains that core aspects of every society are organised around the utilisation of resources (production) as well as maintaining the social system (reproduction). Cultural development results from changes in these two areas. However, even representatives of this approach are very clear-eyed about the fact that the scientific language used in cultural materialism and the terminology applied to describe culture come from an outsider’s observation (etic) and do not explain the meaning of cultural artefacts, communal activities, narratives, and ideas experienced by the research subjects themselves, the insider’s view (emic). The etic and the emic (i.e. the observed and the experienced) remain critically important concepts of analysis within anthropology to this day. The study of local material and energy flows (Rappaport 2000; Singh et al. 2001; Grünbühel et al. 2003) can be regarded as etic analyses of the “infrastructure”, or resource-use base, of local economies, while

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studies on cultural meaning of resource use (Condominas 1977; Gingrich 2002) would look for an emic view on the paramount reason for resource use, production, and consumption. The idea that the natural environment has a strong impact on the shape, structure, and ideology of culture has a long tradition in anthropology: initially as determinism (i.e. a particular natural environment leads to a specific mould of culture; Service 1962); later as a function of resource use (i.e. the more land/energy/material a society uses, the more it transforms specific dimensions of its cultural “core”; White 1943; Steward 1955/1972; Rappaport 2000); even more recently, environment was regarded merely as a limiting factor to particular cultural developments (population growth, scarcity of resources, adaptation to environmental and climatic conditions; cf. Roth & Grünbühel 2013). This succession of theoretical ideas has brought EA very close to and has become utterly compatible with modern ideas of sustainability and adaptation. What anthropologists have shown is that sustainability in not automatically ingrained in pre- or non-industrial societies; however, successful adaptation to their environments has brought populations – in the absence of a growth paradigm – very close to economic conditions that we would deem sustainable today. So, while in the Euphrates and Tigris delta wheat production has historically failed because of a changing climate and the failure to adapt to new conditions (Tainter 1988; Diamond 1997), rice has been grown successfully in Southeast Asia for the last 3000+ years without huge advancements in technology using labour, animal drought, and the plough – a case of successful adaptation (Bray 1994). “Some societies have proven remarkably stable in that they have reproduced a technology which did not alter their environment irreversibly in ways requiring technical innovation or dramatic social change” (Eriksen 2001: 199). A large factor in the understanding of sustainability has proven to be that in non-industrial societies’ producers oversee the entire production process and usually have control over the required resources, while in modern diversified societies (organic societies, according to Durkheim 1997), the individuals see but a specific, limited aspect

of the entire production process and have no resource control (Marx 1872/2001). Due to the “persistence of subsistence” (Waters 2007), producers can establish a close relation between production and consumption, as most goods produced are consumed by the same person/family/group. In addition, labour in non-industrial societies is not a scarce resource and therefore needs not to be optimised or factored in as cost. Furthermore, as part of a community with reciprocal relationships, sustainable use of the local resource base is usually a communal consideration, particularly if the primary resources (land, water, genetic resources, technology) are communal and not individually held (Wittfogel 1957; Lansing 1991). Institutionally, the tendency towards sustainable production has been explained by introducing the concept of underproduction (Sahlins 1974). In local resource use systems, production is often mitigated by risk-aversity over maximisation of productivity. One aspect of this is that local communities depend on local resources and are, therefore, more adversely impacted by local resource shortages. Additionally, however, maximisation of production also creates inequality, which, in turn, endangers social cohesion. In fact, exceptional wealth creation is often prevented by traditional institutions, which rely on the fact that ideologies must be relevant to all members of society. Reciprocity – of goods, of labour, of experiences – only works well in the absence of inequality.3 Sahlins (1974), following the agrarian economist A. Chayanov, states that households with high productive capacity will likely tend to underproduction in order to equalise across the larger group, while households with high dependency ratios will struggle more to increase productivity (cf. Rössler 1999). Sahlins also redefines what he terms an “affluent society”. Rather than regarding absolute levels of consumption, the amount of time used in subsistence and food-producing activities, as well as control over productive means could be considered as wealth that does not even negatively impact on other members of the society. Conversely, wealth in industrial society necessitates the parallel existence of hunger or relative poverty (Sahlins 1974). Today, EA has largely merged with three research approaches: (i) political ecology; Clemens M. Grünbühel

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(ii) livelihoods research; and (iii) adaptation research. Political ecology, being a transdisciplinary research field, investigates nature–society relations from a resource-user perspective. Questions of access to and inequality of control over resources had been addressed in EA and are now carried forward in political ecology research. Many of the theoretical foundations, as well as comparative ethnographic approaches, are shared across EA, ecological anthropology, and political ecology. Critical voices that have spoken out against the political ecology approach emphasise that the material base of the environment and resource use is actually omitted, and studies are often rendered to a sociological exposition of power relations (Karlsson 2015). The study of livelihoods has long been central to EA. While at the outset, livelihoods were conceptualised as more static, ideotypical moulds of livelihoods, the modern study of livelihoods focuses on diversification, adaptability, and fluidity of livelihood types (Dorward et al. 2012). With the conceptualisation of the sustainable livelihoods approach, the realisation has sunk in that modern-day hunters are not just hunters, “peasants” are not just that, and so on. Rather, they are many things at once. Small-scale farmers would conduct a variety of livelihood activities to make ends meet: they farm for food and income; they extract biological and mineral resources from their immediate environment, such as wild vegetables and mushrooms, game, and salt; they sell their labour in urban markets during the agricultural slack season; and they might maintain a homestay service for urbanites interested in community-based agritourism (Hussein & Nelson 1998). Diversified livelihoods mean that (usually rural) people can also not be easily typified as “subsistence”, “market-orientated”, “formal”, “informal”, and so on. They are everything in parallel, and classical economic types may not apply. Instead, analytical approaches tend towards tailored typologies, which reflect the proportion of livelihood activities along the continuum: subsistence – market-dependent; remote – integrated; resource-controlling – resource-dependent, and so forth (Serrano-Tovar & Giampietro 2014; Rigg et al. 2016). Resulting from livelihoods research and combining it with the study of environmental and climatic change, adaptation research Clemens M. Grünbühel

looks at transformations undergone by local economic systems. Following theories developed in cultural ecology, adaptive capacity is seen as a function of the cultural system to transform resource use and economic relations with the aim of minimising the impact of environmental shocks and increasing resilience to external stresses. In this view, culture promotes humans’ ability to adapt through building knowledge, instilling informational feedback loops, enabling the diffusion of innovations, and institutionally applying humans’ ability for reflection (Steward 1972; Rogers 1983; Adger et al. 2003; Roth & Grünbühel 2013; Brown et al. 2019). While the integration of political ecology and EE has largely been achieved (cf. M’Gonigle 1999), the early integration of the livelihoods approach in EE suffers from a mismatch in scale and the tendency to look at households as resource consumers rather than active, empowered agents (Sneddon 2000). The need for integrating questions of traditional livelihoods and loss of environmental services, however, has been shown by Martinez-Alier early on (2001). Newer work allows for combining analytical frameworks to assess climate adaptation options (e.g. Reed et al. 2013). It can only be assumed that further exploration into dynamic livelihoods, resource decisions at the household level (as opposed to the study of consumerism), livelihood typologies, which include levels of material and energy use, and social adaptation/innovation studies would benefit EE into the future. While the contribution of ecological anthropology to ecological economics has been documented (Røpke 2004; Hornborg 2017), the anthropological study of resource transformation, labour, institutions, and cultural adaptability in EA has not yet been acknowledged to the same extent. Hornborg (2017) remarks that ecological anthropology influences have suffered from the chicken-and-egg debate over which one is more dominant over the other, the material base or the institutional structure of society. Regardless, the question as to how resources are used at the local level (Singh et al. 2001), how decisions are made regarding local resource use (Grünbühel & Williams 2016), how parallel economies exist within the same household (Martin et al. 2013), how institutions manifest and reproduce inequalities (Brady et al. 2016), and how questions of

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sustainability and climate adaptation can be answered by observing communities residing along the subsistence–market continuum still rings relevant for a modern approach to the transdisciplinary art form of ecological economics today. Clemens M. Grünbühel

Notes 1.

As a group, however, several cultures have been known to overexploit locally available resources with the consequence of migration, poverty, and even collapse (Turnbull 1972; Tainter 1988). 2. These investigative areas are matched in EE with studies, e.g., on (ecologically) unequal exchange/extractivism (Alonso-Fernandez & Regueiro-Ferreira 2022); external consumption drivers and local impacts (Røpke 2010); the role of green/ecological infrastructure for rural livelihoods (Naumann et al. 2011; Sigwela et al. 2017); challenging economic orthodoxy through ecological thought (Spash 2020); and integrating ecosystem services in sustainable livelihood approaches (King et al. 2019). 3. This refers to the absence of relative inequality within the local community; in the larger state structure, inequality is definitely present, in the form of a managerial and ruling class (Wittfogel 1957; Scott 1976).

References

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in South and Southeast Asia. Climate and Development 11(5): 383–400. Buzney, C., J. Marcoux. n.d. “Cultural Materialism”. https://​anthropology​.ua​.edu/​ theory/​cultural​-materialism/​ Condominas, G. 1977. We Have Eaten the Forest. The Story of a Montagnard Village in the Central Highlands of Vietnam. Farrar, Strauss and Giroux: New York. Diamond, J.A. 1997. Guns, Germs, and Steel. The Fates of Human Societies. W.W. Norton: New York. Dorward, A., S. Anderson, Y.N. Bernal, E.S. Vera, J. Rushton, J. Pattison, R. Paz. 2012. Hanging in, stepping up and stepping out: livelihood aspirations and strategies of the poor. Development in Practice 19: 240–47. Durkheim, E. 1997 (1933). The Division of Labour in Society. The Free Press: New York. Egbert, H. 2018. Marcel Mauss und ökonomische Theorien. Die Institution Geld. Anthropos 113(2): 690–700. Ellen, R.F. 1990. Trade, environment, and the reproduction of local systems in the Moluccas. In: E.F. Moran (ed.), The Ecosystem Approach in Anthropology. From Concept to Practice. University of Michigan Press: Ann Arbor, 191–227. Eriksen, T.H. 2001 (1995). Small Places, Large Issues. An Introduction to Social and Cultural Anthropology. Pluto Press: London. Foucault, M. 1970. The Order of Things. An Archaeology of the Human Sciences. Tavistock: London. Geertz, C. 1983. Local Knowledge: Further Essays in Interpretative Anthropology. Basic Books: New York. Gingrich, A., E. Mader. 2002. Der Elephant im Garten. Einleitende Bemerkungen. In: M. Gingrich (ed.), Metamorphosen der Natur. Sozialanthroplogische Untersuchungen zum Verhältnis von Weltbild und natürlicher Kultur. Böhlau: Wien, 7–30. Godelier, M. 1975. Modes of production, kinship, and demographic structures. In: M. Bloch (ed.), Marxist Analyses and Social Anthropology. Malaby Press: London, 3–27. Grünbühel, C.M., H. Haberl, H. Schandl, V. Winiwarter. 2003. Socio-economic metabolism and colonization of natural processes in SangSaeng: material and energy flows, land use, and cultural change in northeast Thailand. Human Ecology 31(1): 53–86. Grünbühel, C.M., L.J. Williams. 2016. Risks, resources and reason: understanding smallholder decisions around farming system interventions in Eastern Indonesia. Journal of Agriculture and Rural Development in the Tropics and Subtropics 117(2): 295–308.

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25. Economic system Context

It is common that, while partial rationalities are deployed within each of the existing scientific disciplines, what concerns the delimitation of their object of study and their mutual relationships usually appears as matters of fact and not as the result of explicitly agreed rational decisions. Economics does not escape this general trend: the efforts of economists aimed at obtaining explicit enunciative definitions of the object of their science, or of the role of their scientific tasks, were gradually abandoned as they were considered unattainable and unnecessary (Samuelson, 1948, p. 5; Schumpeter, 1954, p. 597). However, as the idea spread that the object of the economic science was something so extremely diffuse that it could not find a definition in the field of words, National Accounts have been developed to offer increasingly detailed calculated versions of the economic system that are generally accepted by economists. Thus, there is the paradox that the agreement that they could not make explicit in the field of words emerges with force in that other more forceful of facts, when the same idea of economic system embodied in the National Accounts classifications and aggregates is taken as a common framework for discussion, offering a version of the study object of economics that, though minimal, is generally accepted by economists. However, the generally unified criteria informing the construction of such accounting systems do not fall from the sky. Rather, they reflect the categories and approaches presiding over that unity in eclecticism of nuance toward which economic thought has drifted (Naredo, 2015a, Ch. 21). And the numerical representation of economic system informs about the underlying axiomatics that can be expressed not only in ordinary language but also in mathematical language, as we will clarify later. The developments of mathematical logic applied in the theory of knowledge today show that the study object of the existing sciences is not usually specified through explicit definitions since it is implicitly delimited by the structure of the axioms governing them. It is the formulation of this system of axioms, and of the definitions shaping this system in a specific applica-

tion model, which yields details about the study object that are unapproachable through definitions, statements, or explicit enumerations. Therefore, it is not the impossibility of delimiting this study object, nor even less the non-existence of inherent limits to it, that explains why it has not yet been possible to give a generally satisfactory version of this study object. We will see now the conditions that allowed the development and unification of National Account systems that shape the usual notion of economic system accepted by most economists. Then, we will discuss the underlying axiomatics required by the theoretical core governing it together with its consequences and alternatives.

Genesis

The consideration of the economic sphere as a possible independent object of study and the idea that it could constitute a coherent system, subject to specific laws, was based from the very beginning on business accounting practices. The very name “political arithmetic” with which economics debuted attests to its close relationship with the accounting record. The application of business accounting principles to society as a whole is not problematic, if—as Adam Smith, the reputed founder of economics, did—individuals are equated with merchants, and states with mercantile companies or associations of merchants. Once making profit had been taken as the common goal of individuals, companies, and countries, and this goal was considered reducible in the homogeneous universe of pecuniary or exchange values ​​in accordance with the current practice of private accounting, it was only necessary to choose the accounts and balances that were considered most significant to examine the economic process. Unsurprisingly, accounting classifications were tributary to the conceptual apparatus at work in economic theory governed by the absolute metaphor of production and the unified and monetary notion of wealth that were imposed after the post-physiocratic epistemological rupture (see the entry 77 in this encyclopedia, on which the present entry is based). Once the notion of production (of monetary value) was taken as the driving force of the economic system, the production account was constructed, represented by the well-known gross domestic product (GDP) as a balance of “added value” (as a result

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of subtracting the value of what was spent on obtaining certain “goods” and “services,” from the monetary value of their sale). The derived notions of distribution, consumption, and accumulation of wealth led to the subsequent construction of the homonymous accounts But that unity of views on the conceptual apparatus that informed the accounting classifications was a necessary, but not sufficient, condition for the existence of National Accounting. It also took adding to the theoretical consensus the loss of faith in the “invisible hand” of the market and the new information requirements of Keynesian macroeconomic policies for the current National Accounts systems to emerge after World War II. International agencies were engaged in establishing a common methodology of the System of National Accounts (SNA) that led to successive versions (1953, 1968, 1993, 2008) in which the production frontier was expanded with debates and conflicts (Lequillier and Blades, 2014). This evolution of GDP-centered flow accounts eventually included patrimonial assets accounts starting with the 1968 United Nations (UN) SNA and its 1993 and 2008 updates, to which the EU adapted with the European System of Integrated Economic Accounts (ESA 1970) and its subsequent versions (ESA 1995 and ESA 2010). Thus, the usual notion of economic system with its defining categories was codified, which allows us to appreciate its capacity as an instrument of analysis as well as its limitations (United Nations, 2014). The methodology of both the UN SNA and EU SNA (ESA 2010) currently in force includes patrimonial assets accounts that would theoretically allow recording the enrichment of “economic agents” associated with the generation and trade of patrimonial assets, as well as the impoverishment caused by extractivism and the overexploitation of natural resources. However, the prevailing accounting systems do not study in-depth these issues and continue to focus on GDP and its derivatives. In other words, while the accounting systems consider an opening balance sheet (which includes the initial value of real economic assets, e.g., land, real estate, etc.) and a closing balance sheet (which includes their final value), there is no breakdown of the loss of assets due to depletion of mining deposits, overexploitation of aquifers,

and so on, or loss of biodiversity and soil fertility associated with ordinary economic operations. Besides, it also grants the same indiscriminate treatment to the “added value” generated by all the “production activities” considered. Thus, the qualification of “satellite accounts” (proposed in the 1993 SNA) applied to the systems of accounts reporting on specific aspects, such as education, health, or the environment, evidences their subordinate nature to the real “National Accounts planet” built on the GDP. And even the effort made to include natural resources in the economic calculation carried out by the 2012 UN System of Integrated Environmental1 and Economic Accounting (UN, 2014) aims to convert physical assets and ecosystems impacts into monetary accounts being compatible with GDP, which thus continues to dominate the accounting systems. The configuration of the usual notion of economic system exemplifies well the steps marked by the sociology of knowledge to socially construct the elaborations of knowledge that are assumed as “reality.” In the book The Social Construction of Reality (Berger and Luckmann, 1966), this process is broken down into three phases: 1) formulation of a world; 2) objectification of that socially produced world; and 3) generalized acceptance of it as something objective and universal. In the case of the usual idea of economic system, these phases are as follows: 1) formulation of this idea associated with the physical world by the French authors of the 18th century, today called physiocrats; 2) displacement, objectification, and quantification of this idea in the form of monetary aggregates of the National Accounts; and 3) generalized acceptance of it as a reality of flesh and blood, forgetting the reasoning that once justified this creation of the human mind. Consequently, the usual notion of economic system and the current quantitative version of it offered by National Accounts provide a significant example of a firm theoretical core that guides economists’ research without being challenged. This notion of economic system has thus created a system of positivities that keep it safe from any criticism. It can only be challenged from the outside, transcending the conceptual apparatus that shapes it, relativizing that notion, and glimpsing the possibility of formulating José Manuel Naredo

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other notions of economic system. For this, it is essential to clarify the content of the firm theoretical core by discovering the axiomatics that inform it that cannot be detailed here. This firm theoretical core is specified in Naredo (1986a, 1986b, 2015a, pp. 547–600), elaborating and communicating the accounting axiomatic that underlies the usual idea of economic system, which allows it to be relativized and transcended, appreciating its limitations and imagining other notions of system to represent the economic process.

Limitations of the usual notion of economic system

Regarding the limitations that the system thus configured entails, it should be noted that many of them come from the monetary reductionism that results from its one-dimensional nature since it is only representable in monetary units, which means ignoring the other dimensions in which the economic process is reflected. And this reductionism is associated with the isolated nature of the system because it works in the isolated universe of monetary values, making its contact with the physical world conditional to it. The conceptual structure that defines the system gives rise, regarding the so-called real economic objects and their corresponding exchange values, to an isolated subsystem, since real economic objects, with their exchange values, are born and disappear inside the system itself, without the need or possibility for it registering any exchange with its environment. They are born when production operations infuse them with monetary value and disappear when consumption operations extinguish them, although, in physical terms, they existed as natural resources before being valued and continued to exist afterward as residues deprived of monetary value. Considering the above, the accounting record of the economic system is maintained only on that fraction of physical assets that are considered producible, abstracting from the consumption of those that are not or presenting them, at most, as something accidental, unpredictable, and alien to the normal operation of such system and that, therefore, have no place in its accounting scheme. By circumscribing the economic field to that of the producible—and the consumable—one abstracts from the fact that a good part of the so-called production—and consumption— José Manuel Naredo

processes involve destruction or degradation for the use of certain natural resources that had not been, and cannot be on a large scale, produced. And by not recording in their accounts this destruction or degradation, even in the case in which those resources have been appropriated and valued, the National Accounts represent a step backward compared with the private accounts: to conform their accounting scheme to the idea of economic system, the former do not practice the distinction so strictly applied in the latter between the operating results produced by the ordinary activity of the company and the extraordinary results, derived from alienating or consuming part of the company’s assets. Thus, for example, mining companies evaluate the mineral reserves contained in the deposits to ensure that they are not depleted before having amortized the investments necessary for their extraction and benefit. On the other hand, National Accounts give income derived from “extractive industries” a treatment similar to that of other activities, ignoring the limited life of mineral deposits, the same treatment as that given to income derived from agriculture, regardless of the fertility of the soil, consumed by depleting practices. This asymmetry that arises between the close-cycle reasoning of the economic system and that of the ecological cycle, necessarily open to energy and materials on which human life is based, not only emerges in the treatment of mining or agriculture but also in all activities—of production or consumption—that benefit from the environment in which they operate and normally contribute to degrading it (and only on some occasions to improving it). This comparative insufficiency of National Accounts is not, however, easily solvable. Well, it cannot be attributed to statistical difficulties or, much less, to the inabilities of the accountants who designed them, but it is imposed by the limitations that the idea of economic system underlying accounting entails. Having postulated that the economic objects to be registered have production as their sole origin and consumption as their sole destination (more or less deferred in time), the notion of amortization also appears as the only mechanism to avoid the degradation of the assets by economic agents: they must establish monetary reserves that compensate the cost of the wearing of their physical assets to ensure their replacement when they reach

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the end of their useful life. The problem is that many of the patrimonial resources that the physiocrats included in part under the denomination of fund assets, are not renewable or producible, or they are not at the rates at which they are consumed, and therefore cannot be replaced.

From the economic system to a systems’ economy, and from the idolatry of GDP to a taxonomy of profit

The formulation of the axiomatics underlying the usual notion of economic system, assumed by most economists and configuring “normal science” in this field, also encourages us to think about the possibility of transcending it by reformulating the mathematical and supposedly quantitative hard core of economic science, thus proposing a true “paradigm shift,” according to the terminology used by Thomas Kuhn in his classic book The Structure of Scientific Revolutions (1962). If, as far as economic science is concerned, we understand by scientific revolution a change of approaches that comes to affect that firm theoretical core that constitutes the usual notion of economic system, either modifying it or removing it from the central place that it occupies today in this discipline, it is obvious that such a revolution is still far from taking place. Despite everything, I believe that, in opposition to the major premises of the enlightened paradigm that largely remains in force (Naredo, 2015b, pp. 17–44)―from which the usual notions of the political and economic system are nurtured―others are emerging that could be grouped around what I have been calling the eco-integrating approach (Naredo, 2015a) as it defends the principle of integration in a triple sense. In the first place, an integration of knowledge transcends the usually divided approaches and, above all, the notorious divorce between economics and ecology. Second, integration of the human species and nature, remembering that symbiosis is the key to the enrichment of life on Earth, induces to shift the current anthropocentrism toward a new geocentrism. And third, integrating the individual and society implies the deep reconstruction of identities and the recreation of civil society itself. Accordingly, the emergence of the eco-integrating paradigm is

not only a matter of politics and economics, but it would have to necessarily encompass the “three ecologies” to which Félix Guattari (1989) refers—the mental, the social, and that of the physical world to be managed—to integrate, in the words of this author, in a new “ecosophy,” both practical and speculative, ethical-political and aesthetic. The greater analytical and predictive potential of the eco-integrating approach, together with its open, transdisciplinary, multidimensional nature and the greater breadth of its object of study, should also enhance its inclusive nature in the face of the reductionist dogmatism of the ordinary economic approach. For example, in my book Taxonomy of Profit2 (Naredo, 2019), I apply multidimensional approaches to make up for the paradoxical lack of the ordinary economic approach which generates a science of profit that does not classify or rank profit forms since it takes them for granted, indiscriminately encompassing them in GDP or ignoring them when they do not appear in it—as is the case especially in recent times with real estate and stock capital gains, or with the creation of bank and financial money. This leads this approach to lose analytical and predictive capacity and to allow businesses that are both more lucrative but less recommendable, from the ecological and social point of view, to flourish. The more inclusive and flexible character of the new eco-integrating paradigm was already maintained in my aforementioned text on axiomatics (Naredo, 1986b, pp. 33–4) when, after emphasizing “the manifest contradiction between the principles that inspire the functioning of the economic system and those of the ecological system,” I believed in the obligation to underline that this would not be serious if, as has happened in other branches of knowledge, old dogmatisms were cornered to give way to other more modest and flexible approaches: today it is not so much a question of describing the system that was supposed to rule in each of these separate worlds, physical, economic, etc., as of studying the various systems that can represent them in order to choose those most in accordance with the context and the purposes in which their application is framed … In the case of economics, the problem is that the acceptance of this multidimensionality implies breaking with the monopoly that the usual notion of economic system, one-dimension-

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150  Elgar encyclopedia of ecological economics ally anchored to the reductionism of monetary value, has exerted [whose essential features are specified by the axiomatics included in the aforementioned texts]. This is difficult, given the absolute character that has been attributed in economic science to this idea of closed system, which explains the scant concern of economists for the analysis of other systems that fall, from various branches of knowledge, on the resources to be managed and their possible utilitarian purposes, with the consequent divorce between economy and ecology.

Indeed, dogmatisms continue to promote the divorce between economics and ecology and make the belief in THE economic system prevail, as THE Newtonian system was once imposed as a dogma of scientific knowledge capable of explaining THE world system, as the title of the well-known book by Laplace (1796), The System of the World, exemplifies. Curiously, in economics, the same system of the economic world continues to prevail since Adam Smith, when in physics itself there have been scientific revolutions that questioned the mechanistic dogma as the only paradigm of rationality, and other systems have appeared with which to interpret and predict the world—like relativistic physics and quantum physics—leaving classical mechanics as a particular systemic notion that continues to be useful for working with modest speeds and instruments of the macroscopic world to which we are accustomed. As for the star observer, the universe of Ptolemy may be more useful than that of Copernicus, although we know that the Sun, the planets, and the stars do not revolve around the Earth. The eco-integrating paradigm proposes to move, finally, from the dogmatism of that single system—THE economic system—to a systems’ economy, fused with that systems’ biology that is ecology. For this, the National Accounting system focused on GDP should lose the absolute prominence that it has today to allow the “satellite accounts”, which report on the physical and social dimensions related to management, to improve and access the category of “true planets”, while the monetary reflection is broadened and qualified to reflect well the entire taxonomy of profit. It is also needed to shift reflection toward the institutional framework that guides the economic process, with its forms of profit and exchange supported by social conventions as relevant as the forms of property and money. José Manuel Naredo

But the problem is not only that the scientific revolution that we have been promoting in economics for so long is still far from being successful. I see that, at this point, it is still not well identified! Even among critics of the status quo who, more concerned with internal criticisms have assumed the usual notion of economic system, thus neglecting external criticisms. Indeed, a recent book by Clive Spash (2020) shows well the prevailing sea of confusion and calls for a necessary pooling of critical thinking, inaugurating a collection titled Inclusive Economics. This requires clearly defining the conflict between the old and the new paradigm and showing the more inclusive nature of the latter to join ranks around it. In Spash’s book, this point remains obscure, although it shows how the so-called “heterodox (or critical) economics” is largely nurtured by researchers who assume the old paradigm embodied in the usual notion of economic system, thus being only critical of certain policies, trends, or interpretations. That said, it seems clear that we must underline that the pending scientific revolution opens a gap between those of us who are putting into practice the new eco-integrating paradigm and working with open, transdisciplinary, and multidimensional approaches, and those who continue to agree with the hitherto usual notion of THE economic system. As a result of what has been said so far, the neoclassical, Keynesian ... or Marxist currents appear as branches of standard economics, because they subscribe to the hard theoretical core that configures the usual notion of THE economic system built on the absolute metaphor of production and its derivatives, exercising only internal criticism on the way to manage it. Marxism, by focusing exclusively on internal criticisms formulated within the framework of the usual notion of economic system, assuming the basic categories of political economy, has contributed to disseminating them and deactivating external criticisms, given its hegemony over critical currents. I hope that over time old marginal group dogmatisms will decay and the pending scientific revolution in economics will be better identified, generating a more conducive context to make the common pooling proposed by Spash consolidate around the emerging eco-integrating paradigm. Well, let us emphasize that the eco-integrating approach does not aim to replace one reduc-

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tionism for another, but rather to place the reflection of monetary values of the ordinary economic approach in a broader and more enriching context that enhances the internal criticisms formulated within the framework of the usual notion of economic system and even improves the monetary reflection itself. In short, it is a matter of broadening and shifting the economic reflection from THE economic system to a systems’ economy and from GDP to a taxonomy of profit. José Manuel Naredo

Notes

1. The very denomination of “environmental accounts” denotes their dependent nature on the SNA centered on GDP, whose prominence leads to understanding through the concept of “externalities” the relationship between the economic process and the physical “environment” where it takes place, which is composed of “free” or “non-economic goods.” The weight of the dominant economic ideology thus imposes the denomination of statistics and even environmental accounts, instead of speaking, as would be logical, of statistics and accounts of natural resources, of the territory, and of the physical (water, energy, materials, etc.) and monetary flows that make up the metabolism of our societies. It is worth saying that, while this latter procedure lends itself much better to the statistical task by directly and unequivocally identifying the object of study, it would also take away the prominence of the dominant economic ideology underlying the current monetary version offered by the National Accounts. Indeed, it would expand the study object and make room for other dimensions and notions of systems to model and interpret this expanded study object, in addition to allowing the recording of the economic process-derived ecological deterioration. Moreover, by contextualizing the problems from this broader and more integrating vision, the very notion of “environment” disappears because, in earth sciences, references to the notion of “environment” are absent, since it is the logic of resources or study objects that establishes systems encompassing this “environment,” resulting from the short-sightedness of the ordinary economic approach. For example, hydrology studies the hydrological cycle as a whole, from the atmospheric phase introducing water by precipitation, followed by runoff, infiltration, and evaporation, in which it loses quantity, quality, and level until it flows into the seas, and the atmospheric phase of the cycle reproduces the process, without any “environment” appearing in this system.

2.

Taxonomía del lucro, in its original Spanish version.

References

Berger, L. and Luckmann, T., 1966, The Social Construction of Reality: A Treatise on the Sociology of Knowledge, New York, Anchor Books. Guattari, F., 1989, Les trois écologies, Paris, Galiée. Kuhn, T.S., 1962, The Structure of Scientific Revolutions, Chicago, University of Chicago Press. Laplace, P.S., 1796, Exposition du système du monde, 2nd ed., Paris, Imprimerie Crapelet. Lequiller, F. and Blades, D., 2014, Understanding National Accounts, 2nd ed., Paris, OECD Publishing. Naredo, J.M., 1986a, “L’axiomatique de l’enregistrement comptable du système économique et les limites de 1’integration d’une comptabilité nationale de patrimoine”, Premier Colloque de Comptabilité Nationale. Etudes de Comptabilité Nationale, E. Archambalt and O. Arkhipoff (eds.), Paris, Económica. Naredo, J.M., 1986b, “La axiomática de la versión usual de sistema económico y sus consecuencias (con especial referencia al tratamiento de los recursos naturales y a la naturaleza de los agregados resultantes)”, Información Comercial Española, 634, http://​elrincondenaredo​.org/​wp​ -content/​uploads/​2020/​10/​Axiom​%C3​%A1tica​ -ICE​-1986​.pdf Naredo, J.M., 2015a, La economía en evolución. Historia y perspectivas de las categorías básicas del pensamiento económico, Madrid, Siglo XXI de España. Naredo, J.M., 2015b, Economía, Poder y Política, Madrid, Díaz&Pons. Naredo, J.M., 2019, Taxonomía del lucro, Madrid, Siglo XXI de España. Samuelson, P.H., 1948, Economics, An Introductory Analysis, New York, McGraw-Hill. Schumpeter, J.A., 1954, History of Economic Analysis, New York, Oxford University Press. Spash, C., 2020, Fundamentos para una economía ecológica y social, Madrid, FUHEM. United Nations, 2014, System of Environmental-Economic Accounting 2012, Central Framework, New York, United Nations.

José Manuel Naredo

26. Economy as an open system Introduction

In general, a system is a collection of interrelated components that, taken as a unified whole, exhibits (emergent) properties that its components do not (Bunge 1979). A general systems theory has been developed since the mid-20th century to explore the basic properties of systems of all types (von Bertalanffy 1968; Boulding 1956). It is possible to consider human society as a broad and complex system comprised of several subsystems that allow it to operate. The existence of economic systems, such as those that provide goods and services vital to the society’s survival, can be empirically distinguished among these subsystems. From an ontological point of view (level of reality), the functioning of the economic subsystem is clearly connected to the rest of the social, natural, cultural, and institutional systems that determine economic results. As Chick (2004: 6) suggests: It is quite clear that not only are all parts of the economic system interconnected to a greater or lesser degree but that the economic system is embedded in and connected with politics, philosophy, history, values, all the elements of social life. Ontologically, then, the economy is unequivocally an open system.

The mainstream neoclassical economics view, however, describes and explains the economic process as a closed system, where economic agents (Homo economicus) and their decisions are largely isolated from external influences and the behavior of other systems. Nonetheless, there has been a long tradition of critique of this neoclassical representation of the economic process (isolated from environmental, social, and cultural factors), pointing to the need to think of the economy as an open system (Kapp 1976; Adkisson 2009). Before we proceed to the central part of this entry, it should be noted that a recent methodological debate has contributed to clarify the notion of open systems (Chick 2004; Chick and Dow 2005; Lawson 1997, 2003; Mearman 2006; Fleetwood 2017). Critical realism’s approach (Lawson 1997,

2003; Fleetwood 2017) defines open systems as those “parts of the social world” that are not “characterized by (stochastic and/or probabilistically specified) regularities between events or states of affairs”—which are considered closed systems (Fleetwood 2017: 41). Other authors, however, have criticized this definition as problematic, since they point out that closed and open systems are not in fact monolithic, and that open systems allow for different degrees of openness (Chick 2004; Chick and Dow 2005; Mearman 2006). In fact, a system can be opened up to environmental, cultural, and institutional factors, as well as to certain scientific disciplines like sociology, psychology, ecology, and others (Arena et al. 2009).

Ecological economics as an approach for analyzing the economy as an open system

From all the possible approaches to representing the economy as an open system, we focus on ecological economics because it also incorporates key elements of other approaches. According to ecological economics, we can also view the economy as an open system from different criteria and perspectives (Carpintero 2013). The best way to gauge the openness of ecological economics might be to compare its view of economic processes with mainstream neoclassical economics. The latter approach treats the economic production process as an isolated system, independent on social and environmental factors. On this carousel of production and consumption, the objects of established monetary value are mechanically moved around. Here, production factors are appropriately valued and, “without loss or friction,” turned into goods and services (Schumpeter 1934: 10) supplied by a circular flow of income from corporations to households and back again. An economic process defined in this way creates an environment where many things are overlooked. Many natural substances and ecosystem functions (photosynthesis, natural purification, biodegradation, or solar energy) constitute preexisting elements that are not assigned market value, or are not producible. In mainstream neoclassical economics, the notion of economic production as the mere “creation of added value” is

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independent from the environmental foundations upon which the system is built. Waste and pollution from industrial processes are also excluded from this definition and are referred to as “externalities.” In any text on standard economic analysis, whether introductory or advanced, the production of goods is described as depending only on two production factors: labor and capital.1 The availability of natural resources is not considered a constraint. In the absence of biophysical links, it is assumed that economic growth is measured solely in terms of monetary value and is based on compound interest, for which there is almost no correlation with the functioning of the physical world. According to the conventional approach, economic growth can simply be defined as the increase of monetary aggregates (at constant prices), such as national income or gross domestic product (GDP), which, by definition, are severely deficient in terms of environmental sustainability. In this case, there is no need to worry about the distribution of resources if natural resources no longer represent a constraint to production, as wealth can be increased limitlessly through economic growth. When the pie size increases, everyone gets a slice without cutting into other people’s slices, the Pareto principle is satisfied as no one is worse off. After environmental issues became so large that the public could no longer avoid them, the mainstream neoclassical economic paradigm attempted a new twist on the old concept. Unfortunately, rather than integrating forms of reasoning and management that go beyond monetary valuation, the conventional economic approach has spent much of its intellectual energy seeking ways to expand “the measuring rod of money” to areas that had historically been excluded from economics. This is due to the lack of markets, which makes it challenging to assign a monetary value. Here we find a paradoxical choice: orthodoxy seeks to solve the problem of monetary valuation by maintaining a conception of the economic system as an isolated system and by enforcing a monetary valuation of areas previously excluded. In terms of the relationship between economy and nature, mainstream neoclassical economics has generally accepted the idea that the environment should be treated as a separate variable in the eco-

nomic system, similarly to labor or capital, for example. Then, the standard toolbox can be applied to it in a conventional manner. Nature and ecosystems are considered subsystems of the economy, but only if they are monetized, rather than seeing the economy as embedded within the larger systems of society and the biosphere. Ecological economics takes the more realistic latter view: the economy is a subsystem embedded in a larger system, the biosphere, and therefore its dynamics are restricted to, and should be compatible with, the laws governing the functioning of the biosphere itself (the laws of thermodynamics and ecology; e.g., Costanza et al. 2015). According to thermodynamics, the economy and society are “open” systems that can exchange both matter and energy with their surroundings, whereas the biosphere is a “closed” system that can exchange only energy with its surroundings, not matter. Unlike the conventional approach, ecological economics considers the production of goods and services as an open system, co-evolving in close collaboration with other social and natural systems (Norgaard 1984; Gowdy 1994). As Gowdy and Erickson (2005: 219) claim, ecological economics “is the school of thought that explicitly recognises the interconnections and interdependencies of the economic, biophysical and social worlds.” In this sense, understanding the economy as an open system involves at least the following issues.2 First, it implies rethinking the representation of the economic process, going beyond the traditional circular flow of income between households and firms to incorporate natural resources before they are assigned a monetary value and waste once they have lost their value. From this perspective, the economic process can be defined as social metabolism: as any organism extracts nutrients from the natural world to live, grow, and reproduce, the economy extracts raw materials, energy, and labor to transform them into consumer goods, infrastructures, and waste (e.g., Ayres and Simonis 1994; Fischer-Kowalski and Haberl 2015). Gowdy and Erickson (2005: 208) suggest that ecological economics may therefore offer a “structuralist approach that categorizes consumption and production as a unified social and biophysical process.” Óscar Carpintero and Jaime Nieto

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Source:

Adapted from Passet (1996: 45).

Figure 26.1 The economy as an open and nested system

Second, from a methodological point of view, and given the scope of “the relationships between ecosystems and economic systems in the broadest sense” (Costanza 1989: 1), the economists must be willing to adopt a transdisciplinary approach with other fields. There are many problems, difficulties, and theoretical or political challenges that occur at the boundaries of established disciplines that make transdisciplinarity essential. This entails building bridges between economics and other social (psychology and sociology) and natural (thermodynamics and ecology) sciences. An ecological economics perspective suggests that it is no longer viable to conduct production and consumption research that does not rigorously integrate behavior and socio-environmental data from disciplines that analyze them. Economists should take into account, for example, lessons from psychology on shortcomings associated with the “canonical model of economic behaviour” (Homo economicus; e.g., Simon 1987; Gowdy 2008) that can be opposed by alternative frameworks such as the bounded rationality approach (Simon 1997); the principles of subordination, satiation, and lexicographical order put forward Óscar Carpintero and Jaime Nieto

by Georgescu-Roegen (1954) or the Political Economic Person, highlighting the role of ideology in individual behavior (Söderbaum 2018); or the teachings of thermodynamics regarding the limitations of energy, materials, and wastes (Georgescu-Roegen 1971; Ruth 1993; Baumgärtner et al. 2007; Valero and Valero 2014) as they affect economic management, and the proper measurement units (biophysical or land) for analyzing these issues. A wider range of options should complement monetary and financial measurements that focus on a single (monetary) unit. As a practical matter, it requires going beyond evaluation and decision-making tools such as cost–benefit analyses, which are based on a single monetary criterion of accounting and choice, and introducing rigorous multi-criteria approaches that encompass a variety of dimensions (Munda 2008; Brown et al. 2017). Third, and precisely for that reason, ecological economics is skeptical of mainstream strategies for analyzing environmental sustainability in a closed system, using single-unit measurements or “ecologically reformed” monetary indicators such as Green GDP or genuine savings to assess the environmental

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sustainability of a country or region (the “weak sustainability” approach; Pearce and Atkinson 1993; Neumayer 2010; World Bank 2022). Ecological economics upholds the “strong sustainability” approach and insists that the economy must be seen as a subsystem of the environment, thus sustainability should be analyzed as a “scale issue” (Daly 1992). It refers to the size of the economic system within the biosphere, which raises an important question: How can this size be measured? Ecological economists proposed at least two alternatives: (1) a “resource flows approach” based on accounting for energy and materials that flow through the economy (socioeconomic metabolism); and (2) a “land use approach” based on estimating how much territory is needed for the production and consumption of a country (e.g., ecological footprint, land use–land cover approach). Both alternatives allow us to obtain information about the capacity of ecosystems to supply resources and to absorb waste and emissions. Ecological economics, as an open system approach, is therefore concerned with highlighting the limitations of the consumption of both energy and material as constraints on any situation of unrestricted growth in the production of goods and services. A fourth issue emphasized in the open systems approach of ecological economics has to do with analyzing the rules of the game for economic activity. Markets are economic institutions that cannot operate or function without rules governing their functioning, and those rules are generally established by the public sector. These rules of the game determine and guarantee property rights, define costs, penalties, incentives, labor conditions, and the requirements for the development of production processes. Thus, “efficiency” (which relates production and costs), “profitability,” or “optimum” results are not parameters that appear out of thin air, but are rather influenced both by the institutional framework that regulates and defines the market itself and by the distribution of income and wealth (economic power). From this viewpoint, a bridge has been built between ecological economics and the teachings of old institutional economics applied to natural resources. A large part of this work was done by pioneers in the mid-20th century, such as S. Ciriacy-Wantrup (1952) or K.W. Kapp (1950), who understood the

relevance of some of the findings of the old institutionalism—namely, Veblen and Commons—for the analysis of the relationships between the economy and nature. Ciriacy-Wantrup and Kapp have been recovered as a link by the modern ecological version of the “old” institutionalism (Bromley 1991; Vatn 2005; Ostrom 1990; Berger 2008; Söderbaum 2021). Along these lines, there are three long-standing, foundational questions at the base of any institutional reflection on the open relationship between the economy and natural systems, and about management and the use of natural resources, which have been skillfully addressed (Bromley 1982: 842–3): a) Who controls the rules of management (institutions) that determine the rate of use of natural resources?, b) Who is in a position to receive the benefits of a certain criterion of use?, and c) Who must pay the costs associated with the use of natural resources? To answer these questions, we must look beyond the usual mechanisms of market exchanges and decide to focus on the processes of economic and social power, the concentration of income and wealth, and how they are distributed, as these processes are independent and determine these mechanisms. We have long known that every economic decision is preceded by a distributive conflict, and the development of any activity has implications for distribution. From this perspective, environmental problems are largely associated with political dimensions of unequal appropriation of environmental space and decision-making. Hence, reversing the process of ecological degradation will be difficult without critically analyzing the rules of the game and power structures that influence economic performance. Scientifically, shaping all of this is not a simple task. In a seminal article about the economy as an open system, Kapp (1976: 97) pointed out more than four decades ago: We need a new approach which makes it possible to deal with the dynamic interrelations between economic systems and the whole network of physical and social systems and, indeed, the entire composite system of structural relationships. It would be an illusion to believe that such a system view of the economy can or will emerge from the traditional modes of analytical thinking … Systems thinking is inevitably complex inasmuch as it is concerned

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156  Elgar encyclopedia of ecological economics with discontinuous non-linear “feedback” effects which characterise the dynamic interdependencies between the different systems as well as of each subsystem with the composite whole.

Not surprisingly, Kapp’s reflection was made shortly after one of the first attempts to think of the interrelationships of the economic system and natural systems on a planetary scale, applying a powerful methodology such as system dynamics (Meadows et al. 1972). Undoubtedly, this approach specialized in the treatment of “discontinuous non-linear feedback effects,” and “dynamic interdependence” should become a common tool for all those ecological economists and natural scientists who want to think of the economic system as an open system and thus draw the appropriate conclusions from it (Capellán-Pérez et al. 2020; D’Alessandro et al. 2020). By viewing the economy as an open system, a powerful scientific metaphor is created. In other words, it means embracing the fact that what we refer to as the economic system is embedded in the social system and in the biosphere. This opens up exciting possibilities for scientific reflection and well-founded policy making. Óscar Carpintero and Jaime Nieto

Notes

1. Classical influence is seen in the tendency to include “land” nominally as a third factor of production in economic textbooks. However, we should not be misled on this point. It is easily observable that these textbooks and all subsequent developments related to production functions in micro- and macroeconomics generally avoid actually referring to land as a production factor and focus only on labor and capital, turning land into a merely nominal factor. 2. Many of them were already apparent in the pioneering and transdisciplinary works of Georgescu-Roegen (1971), Boulding (1956), Daly (1968), Kapp (1950), Ciriacy-Wantrup (1952), Martínez Alier (1987), and Naredo (1987).

References

Adkisson, R.V., 2009. The Economy as an Open System: An Institutionalist Framework for Economic Development, in: Tara Natarajan, Wolfram Elsner, Scott Fullwiler (eds.), Institutional Analysis and Praxis, Dordrecht: Springer, pp. 25–38.

Óscar Carpintero and Jaime Nieto

Arena, R., Dow, S., Klaes, M., eds., 2009. Open Economics: Economics in Relation to Other Disciplines. London: Routledge. Ayres, R.U., Simonis. U., eds., 1994. Industrial Metabolism: Restructuring for Sustainable Development. New York: United Nations University Press. Baumgärtner, S., Faber, M., Schiller, J., 2007. Joint Production and Responsibility in Ecological Economics. On the Foundations of Environmental Policy. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Berger, S., 2008. K. William Kapp’s Theory of Social Costs and Environmental Policy: Towards Political Ecological Economics. Ecological Economics, 67(2), 244–52. Boulding, K.E, 1956. General System Theory, The Skeleton of a Science. Management Science, 2, 197–208. Bromley, D.W., 1982. Land and Water Problems: An Institutional Perspective. Journal of American Agricultural Economics, 64(5), 834–44. Bromley, D.W., 1991. Environment and Economy. Oxford: Basil Blackwell. Brown, J., Söderbaum, P., Dereniowska, M., 2017. Positional Analysis for Sustainable Development. Reconsidering Policy, Economics and Accounting. London: Routledge. Bunge, M., 1979. Treatise on Basic Philosophy. A World of Systems. Dordrecht and Boston: Reidel. Capellán-Pérez, I., de Blas, I., Nieto, J., de Castro, C., Miguel, L.J., Carpintero, Ó., Mediavilla, M. . . . , 2020. MEDEAS: A New Modeling Framework Integrating Global Biophysical and Socioeconomic Constraints. Energy Environmental Science, 13, 986–1017. Carpintero, Ó., 2013. When Heterodoxy Becomes Orthodoxy: Ecological Economics in The New Palgrave Dictionary of Economics. American Journal of Economics and Sociology, 72, 1287–1314. Chick, V., 2004. On open systems. Brazilian Review of Political Economy, 24, 1–16. Chick, V., Dow, S., 2005. The meaning of open systems. Journal of Economic Methodology, 12(3), 363–81 Ciriacy-Wantrup, S.V., 1952. Resource Conservation: Economics and Policies. Berkeley: University of California Press. Costanza, R., 1989. What is Ecological Economics? Ecological Economics, 1, 1–7. Costanza, R., Cumberland, J.H., Daly, H., Goodland, R., Norgaard, R.B., Kubiszewski, I., Franco, C., 2015. An Introduction to Ecological Economics. Boston: CRC Press. D’Alessandro, S., Cieplinski, A., Distefano, T., Dittmer, K., 2020. Feasible Alternatives to Green Growth. Nature Sustainability, 3, 329–35.

Economy as an open system  157 Daly, H.E., 1968. On Economics as a Life Science. Journal of Political Economy, 76, 392–406. Daly, H.E., 1992. Allocation, Distribution, and Scale: Towards an Economics that is Efficient, Just, and Sustainable. Ecological Economics, 6, 185–93. Fischer-Kowalski, M., Haberl, H., 2015. Social Metabolism: A Metric for Biophysical Growth and Degrowth, in: J. Martínez Alier, R. Muradian (eds.), Handbook of Ecological Economics, Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing, pp. 100–138. Fleetwood, S., 2017. The Critical Realist Conception of Open and Closed Systems. Journal of Economic Methodology, 24, 41–60. Georgescu-Roegen, N., 1954. Choice, Expectations and Measurability. The Quarterly Journal of Economics, 68(4), 503. https://​doi​ .org/​10​.2307/​1881875 Georgescu-Roegen, N., 1971. The Entropy Law and the Economic Process. Cambridge, MA: Harvard University Press. Gowdy, J., 1994. Coevolutionary Economics: The Economy, Society and the Environment. Dordrecht: Kluwer Academic Publishers. Gowdy, J., 2008. Behavioral Economics and Climate Change Policy. Journal of Economic Behavior and Organization, 68(3–4), 632–44. Gowdy, J., Erickson, J., 2005. The Approach of Ecological Economics. Cambridge Journal of Economics, 29, 207–22. Kapp, K.W., 1950. The Social Costs of Private Enterprise. Cambridge, MA: Harvard University Press. Kapp, K.W., 1976. The Open-System Character of the Economy and its Implications, in: K. Dopfer (ed.), Economics in the Future, Boulder, CO: Westview, pp. 90–105. Lawson, T., 1997. Economics and Reality. London: Routledge. Lawson, T., 2003. Reorienting Economics. London: Routledge. Martínez Alier, J., 1987. Ecological Economics: Energy, Environment and Society. Oxford: Basil Blackwell. Meadows, D.H., Meadows, D.L., Randers, J., Behrens III, W.W. 1972. The Limits to Growth; a Report for the Club of Rome's Project on the Predicament of Mankind. New York: Universe Books. Mearman, A., 2006. Critical Realism in Economics and Open-Systems Ontology: A Critique. Review of Social Economy, 64, 47–75. Munda, G., 2008. Social Multi-Criteria Evaluation

for Sustainable Development. Heidelberg: Springer Verlag. Naredo, J.M., 1987. La Economía en Evolución. Madrid: Siglo XXI. Neumayer, E., 2010. Weak Versus Strong Sustainability: Exploring the Limits of Two Opposing Paradigms. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Norgaard, R.B., 1984. Coevolutionary Development Potential. Land Economics, 60, 160–73. Ostrom, E., 1990. Governing the Commons: The Evolution of Institutions for Collective Action. New York: Cambridge University Press. Passet, R. 1996. Principios de Bioeconomía. Madrid: Fundación Argentaria-Visor Distribuciones. (Spanish translation of L’Economique et le Vivant, Paris: Payot, 1979). Pearce, D.W., Atkinson, G.D. 1993. Capital theory and the measurement of sustainable development: an indicator of “weak” sustainability. Ecological Economics, 8 (2), 103–08. Ruth, M., 1993. Integrating Economics, Ecology and Thermodynamics. Dortrecht: Kluwer Academic Publishers. Schumpeter, J., 1934. The Theory of Economic Development. Cambridge, MA: Harvard University Press. Simon, H. A., 1987. Behavioral Economics, in: The New Palgrave Dictionary of Economics, Vol. 1, J. Eatwell, M. Mulgate, P. Newman (eds.), London: Macmillan, pp. 221–4. Simon, H. A., 1997. Models of Bounded Rationality: Empirically Grounded Reason (Vol. 3). Cambridge, MA: MIT Press. Söderbaum, P., 2018. Economics, Ideological Orientation and Democracy for Sustainable Development (2nd ed.). Bristol: World Economics Association (WEA) Books. Söderbaum, P., 2021. The Challenge of Sustainable Development: From Technocracy to Democracy-Oriented Political Economics. Economic Thought, 10(1), 1–13. Valero, A., Valero, A., 2014. Thanatia: The Destiny of the Earth’s Mineral Resources. London: World Scientific Publishing. Vatn, A., 2005. Institutions and Environment. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. von Bertalanffy, L., 1968. General System Theory. New York: George Braziller. World Bank. 2022. World Development Indicators. https://​www​.worldbank​.org

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27. Ecosystem services Humankind has learned to insert itself into all the biological communities of the planet. During the last 200 000 years, Homo sapiens, a very recent species in the world, has colonized and become natural in every single corner of the Earth (Boivin et al., 2017). In less than 3 million years since we made our appearance in the plains of Eastern Africa, we have invaded the whole surface in every continent and, thanks to our unique abilities to read and use each type of ecosystem, we have inserted ourselves into its trophic webs, with different levels of effort and success (Roberts and Stewart, 2018). The process that has allowed us to learn and shape our ecological identity in every place is recognized as culture, the sum of symbolic and material devices that we have created to thrive through time and travel (Henrich, 2011). Now, the human population has reached a peak of nearly 8 billion, despite internal competition and maladaptation, and will probably stabilize in a few decades (Vollset et al., 2020), albeit any big surprise. Ecosystems are complex in nature and have evolved in a very plastic way over millions of years to allow adaptation to the physical changes of the environment as well as to the dynamics of living systems: biological diversity has evolved as the result of its own growing complexity, gaining capacities to endure; they have acquired resilience (Gunderson, 2000). Competition and cooperation among species have shaped titillating communities of beings, permanently exchanging energy to survive and reproduce, a billion years’ functional network where we, humans, were able to insert ourselves through imagination and skills development (Haviland et al., 2013). We now call “ecosystem services” the amount of matter, energy, and information that flows into our bodies and immense population (Daily, 1997), and our “ecological footprint” is the amount of matter, energy, and information we put back into the non-human world (Wackernagel and Beyers, 2017). Ecosystem services can be depicted as the transference of the physical content (biomass) of non-human living communities to humans, as well as the capture by us (Homo sapiens) of the biological processes (functions) that link all beings. By doing this, certain amounts

of matter and energy are appropriated by societies with different levels of awareness, tools, and behavioral modifications (Jacobs et al., 2013). The process creates, in parallel, some capacities to assess the level of engagement of humans in what may be called their environment, and through that, pushes the understanding of the effects of their activity in different scales of time and space; as a result, a human population may thrive or collapse (Constanza et al., 2007). Ecosystem services, then, is an expression that covers all the flows from the physical and biological environment to the humankind, but also the induced mental processes. Therefore, we may speak of tangible or non-tangible ecosystem services: capturing and eating a fish brings food and happiness to a person, creating satisfaction for the body at the same time as an ephemeral sense of well-being, felt as emotion and able to be recorded, in many ways, into our brains, and shared with other people through language and other symbolic devices. Imagination and communication, thus, are a less tangible result of our insertion in food web networks, but an equally important issue for our survival, as well as an energy-demanding process. Dreams are the result of the human body and mind entanglement with the ecosystem, and can be understood as the way we, as a species, behave and create a coherent image of the world to survive, an issue studied by the emerging field of neuroecology (Riffell and Rowe, 2016). Arts, then, are considered a result of the inspiration we get from our daily experience in the living community, and a critical way to inquire about our deepest embodiment of the world (Milcu et al., 2013). The word “service,” attached to any kind of energy or matter flow, represents the contribution of the ecosystem to human well-being, but it also has many negative meanings: parasites undermining human health, botanical poisons, or hurricanes and floods are all common issues around humans that come with our insertion into the myriad relationships between microbes, animals, plants, and the physical environment where we live, including our built environment (Remme et al., 2021). Humans have created technology and modified their behavior also with impressive diversity, not just to adapt to the material conditions of ecosystems, but to give coherence to the mental models or worldviews necessary to adapt to the multiple scales

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Source:

Pascual et al. (2017).

Figure 27.1 Diverse values related to nature, nature’s contributions to people, and a good quality of life

where living beings interact: we may be aware that the Ebola virus may jump into our bodies if we get close enough to the monkeys inhabited by it, but still become invaded and killed by the smallest particle of living matter we know of, a chunk of lethal RNA. At the same time, glaciation cycles, probably linked to millions of years of solar activity, have changed the planet in ways we have to imagine, since evidence of that in our lifespan is difficult to grasp (Vincent et al., 2004). Valuing ecosystem services has been a complex task recently attempted by some societies (Dendoncker et al., 2013). A larger

interpretation of the importance of the ecological contributions to our well-being are being built to avoid the reduction of complex processes to a single indicator such as monetary value, although economic valuation can contribute to improving decision-making at some point in the modern world (Dasgupta, 2019). Figure 27.1 depicts the effort from Pascual et al. (2017) to summarize ecosystem services under the larger umbrella of “nature’s contributions to people,” arguing that there are many ways to understand how humans capture and value non-human induced ecological processes. The Intergovernmental Brigitte L.G. Baptiste

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Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES) released its “values assessment” (Balvanera et al., 2022) with a summary of the methods currently used to value those contributions, considering ecosystem services as a way to understand them under certain cultural and institutional circumstances. The core concept of nature, however, makes it very difficult to build a common understanding of the ecological flows being captured by humans and their consequences, since many schools of thought, as well as other knowledge systems, do not have a separate notion of non-human beings acting in the world (Ducarme and Couvet, 2020). Since technology and technological devices have been part of our adaptive path through history, and have produced major changes on ecosystems, the resulting modified environments also produce changes in the flows of matter, energy, and information. Those changes may be felt differently according to connectivity, affecting the distribution of ecosystem services through time and space and, therefore, human communities (Englund et al., 2017). Connectivity, a feature of any system that defines pathways for matter, energy, and information, can be and has been modified by human policy at some point, but most ecological processes have their own rationale: biogeochemical cycles, such as those from atmospheric gases, have not been controlled by humans, creating a new planetary reality with potential dire effects, such as global warming (Pachauri and Meyer, 2014; Intergovernmental Panel on Climate Change, 2022). Other massive ecosystem changes have been produced by humans by polluting rivers and oceans with plastics and new substances, overharvesting animal populations, or replacing diversity and complex processes by simpler ones, a process affecting the “safe operating space for humanity” (Rockström et al., 2009). The effects of ecological changes on the flow of services is huge, and humanity struggles to understand its own role in the systems being affected, as well as the differential effects among human and non-human populations. The distributional effects of capturing, using, or transforming ecosystem capacities is being permanently assessed by institutions since we are still competing for clean water, nutritional assets, or friendlier landscapes to live in (Martínez-Alier and Roca-Jusmet, Brigitte L.G. Baptiste

1992). The different and asymmetric levels of capture of any amount of matter or energy by a human population creates a basic economic dilemma that has fueled a new understanding of ecological transformations and its relation to justice (Martínez-Alier et al., 2016), since change produced by humans has thresholds and has a clear effect on other humans. Ecological services distribution is a governance challenge for every society (Constanza et al., 2019). Inquiry on the nature of such effects or consequences is fundamental to create a balanced way of living in a planet that is not the same as the one we evolved in, but where we, as modern humans, neither are the same. Brigitte L.G. Baptiste

References

Balvanera, P., U. Pascual, M. Christie, B. Baptiste, and D. González-Jiménez. (eds). 2022. Methodological Assessment of the Diverse Values and Valuation of Nature of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services. IPBES Secretariat. https://​doi​.org/​10​.5281/​zenodo​ .6522522 Boivin, N., R. Crassard, and M. Petraglia. 2017. Human Dispersal and Species Movement: From Prehistory to the Present. Cambridge University Press. Costanza, R., L. Graumlich, and W. Steffen. 2007. Sustainability or Collapse? An Integrated History and Future of People on Earth. MIT Press. Costanza, R., B. Low, J. Wilson, and E. Ostrom. 2019. Institutions, Ecosystems, and Sustainability. CRC Press. Daily, G. 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. Island Press. Dasgupta, P. 2019. The Economics of Biodiversity: The Dasgupta Review. HM Treasury. Dendoncker, N., H. Keune, S. Jacobs, and E. Gómez-Baggethun. 2013. Inclusive Ecosystem Services Valuation. In: S. Jacobs, N. Dendoncker, and Hans Keune (eds), Ecosystem Services, Elsevier, 3–12. https://​doi​.org/​10​ .1016/​B978​-0​-12​-419964​-4​.00001​-9 Ducarme, F., and D. Couvet. 2020. What Does ‘Nature’ Mean? Palgrave Communications, 6, 14. https://​doi​.org/​10​.1057/​s41599​-020​-0390​-y Englund, O., G. Berndes, and C. Cederberg. 2017. How to Analyse Ecosystem Services in Landscapes—A Systematic Review. Ecological Indicators, 73, 492–504. https://​doi​.org/​10​ .1016/​j​.ecolind​.2016​.10​.009 Gunderson, L.H. 2000. Ecological Resilience— In Theory and Application. Annual Review

Ecosystem services  161 of Ecology and Systematics, 31(1), 425–39. https://​doi​.org/​10​.1146/​annurev​.ecolsys​.31​.1​ .425 Haviland, W.A., H. Prins, and B. McBride. 2013. Anthropology: The Human Challenge. The Thompson Corporation. Henrich, J. 2011. A Cultural Species: How Culture Drove Human Evolution. A Multi-Disciplinary Framework for Understanding Culture, Cognition and Behavior. Psychological Science Agenda. Intergovernmental Panel on Climate Change (IPCC). 2022. Climate Change 2022: Contributions of Working Groups I, II, and III to the Sixth Assessment Report of the Intergovernmental Panel on Climate Change. IPCC. Jacobs, S., N. Dendoncker, and ‎H. Keune. 2013. Ecosystem Services: Global Issues, Local Practices. Elsevier. Martínez-Alier, J., and J. Roca-Jusmet. 1992. Economía Ecológica y Política Ambiental. Fondo de Cultura Económica. Martínez‐Alier, J., L. Temper, D. Del Bene, and Arnim Scheidel. 2016. Is there a Global Environmental Justice Movement? Colloquium Paper No. 16 at: Global Governance/Politics, Climate Justice & Agrarian/Social Justice: Linkages and Challenges. International Institute of Social Studies. Milcu, A.I., J. Hanspach, D. Abson, and J. Fischer 2013. Cultural Ecosystem Services: A Literature Review and Prospects for Future Research. Ecology and Society, 18(3), 44. http://​dx​.doi​ .org/​10​.5751/​ES​-05790​-180344 Pachauri, R.K., and L.A. Meyer. (eds). 2014. Climate Change 2014: Synthesis Report. Contribution of Working Groups I, II, and III to the Fifth Assessment Report of the IPCC. IPCC. Pascual, U., P. Balvanera, S. Díaz, G. Pataki, E. Roth, M. Stenseke, R.T. Watson, et al. 2017. Valuing Nature’s Contributions to People:

The IPBES Approach. Current Opinion in Environmental Sustainability, 26–27, 7–16. https://​doi​.org/​10​.1016/​j​.cosust​.2016​.12​.006 Remme P.R, H. Frumkin, A.D. Guerry, A.C. King, L. Mandle, Ch. Sarabu, G.N. Bratman, et al. 2021. An Ecosystem Service Perspective on Urban Nature, Physical Activity, and Health. PNAS, 118(22), e2018472118. https://​doi​.org/​ 10​.1073/​pnas​.2018472118 Riffell, J., and A.H. Rowe. 2016. Neuroecology: Neural Mechanisms of Sensory and Motor Processes that Mediate Ecologically Relevant Behaviors: An Introduction to the Symposium. Integrative and Comparative Biology, 56(5), 853–5. https://​doi​.org/​10​.1093/​icb/​icw109 Roberts, P., and B.A. Stewart. 2018. Defining the ‘Generalist Specialist’ Niche for Pleistocene Homo Sapiens. Nature Human Behaviour, 2, 542–50. https://​doi​.org/​10​.1038/​s41562​-018​ -0394​-4 Rockström, J., W. Steffen, K. Noone, Å. Persson, F. Stuart Chapin III, E.F. Lambin, T.M. Lenton, et al. 2009. A Safe Operating Space for Humanity. Nature, 461, 472–5. https://​doi​.org/​ 10​.1038/​461472a Vincent, W.F., D. Mueller, P. Van Hove, and C. Howard-Williams. 2004. Glacial Periods on Early Earth and Implications for the Evolution of Life. In: J. Seckbach (ed), Origins. Cellular Origin, Life in Extreme Habitats and Astrobiology, Vol 6., Springer, 481–501. https://​doi​.org/​10​.1007/​1​-4020​-2522​-X​_29 Vollset, E.S., E. Goren, C. Yuan, J. Cao, A.E. Smith, and T. Hsiao. 2020. Fertility, Mortality, Migration, and Population Scenarios for 195 Countries and Territories from 2017 to 2100: A Forecasting Analysis for the Global Burden of Disease Study. The Lancet, 396, 1285–1306. https://​doi​.org/​10​.1016/​S0140​-6736(20)30677​ -2 Wackernagel, M., and B. Beyers. 2017. Ecological Footprint: Managing Our Biocapacity Budget. New Society Publishers.

Brigitte L.G. Baptiste

28. Emergy accounting Introduction

Emergy accounting (EMA; Odum, 1988, 1996; Brown et al., 2016) is an environmental accounting method aimed at assessing the environmental performance and sustainability of processes and systems on the global scale of biosphere and human-dominated systems (Brown et al., 2006; Campbell and Brown, 2012; Lee and Brown, 2021), taking into account free environmental inputs (e.g., solar radiation, wind, rain, and geothermal flows), materials and energy from the outside economy, as well as the indirect environmental support embodied in human labor and services. The emergy definition (the “M” in emergy is sometimes referred to as “energy memory”) is different from the general concept of “energy,” representing the cumulative support that a process or system receives from the biosphere across time and space. This concept has been evolving since the 1950s with H.T. Odum’s (1951, 1957, 1971) work tracing energy flows in ecosystems. With his first attempts at defining emergy, Odum used an energy hierarchy concept as a means of explaining the work of nature and society, concentrating the available energy of resources (exergy, work potential) from lower to higher hierarchical levels (quality increase). When viewed in their wholeness, the systems of nature and society are interconnected in webs of energy flows. Odum’s idea was that all the energy transformations of the geo-biosphere could be arranged in ordered series to form an energy hierarchy, with many joules of sunlight required to make a joule of organic matter, many joules of organic matter to make a joule of fuel, many joules of fuel required to make a joule of electric power, and so on. Within such a “donor-side” perspective, the “value” of a resource relies on the effort that is displayed for its generation by nature and processing by society, over an evolutionary “trial-and-error” process that ensures the optimization of a resource cycle. Mainstream economic theories address the concept of value in monetary terms (willingness to pay, i.e., a user-side concept of value), while the emergy value is related to the amount of primary resources (solar energy, geothermal

heat, etc., over time) invested by nature for resource cycling and generation of life (generation of fossil oil and uptake of carbon dioxide from atmosphere both require the same photosynthetic activity, independently of how much we are willing to pay for an oil barrel). The emergy accounting method is therefore a technique of quantitative evaluation that determines the environmental value (sustainability in terms of biosphere time and space for the process to occur) of non-marketed and marketed resources, services, commodities, and storages in common equivalent units of cumulative solar energy (sej, solar equivalent joules or solar emjoules). In the EMA method, all resource inputs supporting a system are accounted for in terms of their solar emergy (sej), defined as the total amount of solar available energy (exergy) directly or indirectly required to generate a product or support a flow or service (Odum, 1996). In his seminal book, Environmental Accounting: Emergy and Environmental Decision Making, Odum (1996) clearly summarized the theoretical basis of emergy research and provided a complete evaluation procedure for EMA. The first step in EMA is to design an energy diagram of the system (Figure 28.1) where the main areas of energy use, components and their relationships, as well resource inflows and outflows are identified (Brown, 2004). For instance, the diagram in Figure 28.1 can represent an agricultural activity where: R are the “free” renewable resources from nature (sunlight, wind, geothermal heat, etc.), N are the “free” local nonrenewable sources (topsoil, groundwater, biodiversity, etc.), F are the “purchased” imported resources from the larger economy (chemicals, diesel, tractors, etc.), L is the direct human labor, and S represents the indirect labor (services) applied to make F available to the system. Then, the different resource flows crossing the boundaries of the investigated system are organized and quantified in emergy evaluation tables by multiplying their raw amounts by unit emergy values (UEVs, conversion factors; Brown and Ulgiati, 2004). The UEVs quantify the emergy required to obtain a unit (sej/g, sej/J, sej/$, sej/hour) of good or service. When the unit is sej/J, the UEV is most often named transformity; when the unit is sej/g, the UEV is often named specific emergy. In the third step, the emergy flows of the evaluated system are normally aggregated

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according to the above-mentioned specific categories (R, N, F, L, and S). Finally, the total emergy driving the process as well as a series of emergy-based performance indicators (Odum, 1996; Brown and Buranakarn, 2003; Brown and Ulgiati, 2004) can be quantified to support comparisons and discussions about the overall sustainability of one system from an eco-centric perspective.

Emergy indicators

The total emergy converging to a system, U, measures the environmental support provided by the biosphere and needed to produce a process yield Y (output). The total U (= R+N+F+L&S) is a measure of the emergy cost of the yield, that is, the emergy associated to the yield Y (UY). The UEVs measure the global conversion efficiency over the whole chain of processes leading from primary resources to the final product, regardless of whether the driving emergy is renewable or not. UEVs are not sensitive to the renewable-vs.-nonrenewable alternative. When comparing and testing alternative parallel processes (e.g., two power plants generating electricity but running on different energy sources), the UEVs measure their efficiency in delivering the same product (i.e., a kWh of electricity). This kind of efficiency is sometimes named “parallel quality” (Brown and Ulgiati, 2004) and compares systems aimed at the same kind of yield. This calculated efficiency may not be the best possible one because the systems may not have undergone a sufficiently long testing period for improvement and selection. Instead, if a system has undergone a sufficient testing time of its contribution and usefulness to the larger scale in which it is embedded, the UEV indicates its hierarchical position related to other systems in the thermodynamic scale of the biosphere (e.g., photosynthetic, herbivore, and carnivore organisms; agriculture, agro-industry, food distribution; and so on) and can be regarded as a different kind of quality factor (i.e., position, role, feedback control action, relative importance), the so-called “cross quality” (Brown and Ulgiati, 2004). The Emergy Yield Ratio, EYR = U/F, is a measure of the ability of a process to exploit and make available the local renewable and nonrenewable resources by investing outside resources. It provides a look at the

process from a “self-reliance” perspective. The ability to exploit local resources by a process can be read as a potential additional contribution to the economy, generated by investing resources already available (Brown and Ulgiati, 1997). The lowest possible value of the EYR is 1 when the imported emergy used up to generate the yield does not allow any additional exploitation of local resources (U = F), that is, the process is not usefully exploiting any local resource. Processes with EYR equal to 1 or only slightly higher do not provide any significant net emergy to the economy and only transform resources that are already available from previous processes. In so doing, they act as consumer processes more than creating new opportunities to the system. Primary energy sources (crude oil, coal, natural gas, uranium) usually show EYRs greater than 5 since they are exploited by means of a small input from the economy, providing much greater emergy flows, generated by previous geologic and ecosystemic activities over past millennia. Secondary energy sources and primary materials, like cement and steel, show EYRs varying in the range from 2 to 5, indicating moderate reliance on additional local resources. The Environmental Loading Ratio, ELR = (N+F)/R, is designed to compare the amount of locally nonrenewable and imported emergy (N+F) with the amount of locally renewable emergy (R). Considering that an ecosystem in equilibrium with the global biosphere (e.g., a pristine forest, a wilderness protected area) only or mainly uses renewable sources, any increase of the ELR from the zero-reference level translates into a displacement of the system from the environmental equilibrium state. In the absence of human-driven investments (note: subtracted from other areas), the locally available renewable resources would have driven the growth of a mature ecosystem, consistent with the constraints imposed by the environment and characterized by an ELR equal to 0. Instead, the nonrenewable imported emergy drives a different site development. The higher the ELR, the bigger the distance of the development from the fully natural state that could have developed locally without the convergence of nonrenewable investment (either local or from outside). The ELR is therefore a measure of such disturbance to the local environmental dynamics (either in the form of pollution or Silvio Viglia and Sergio Ulgiati

Figure 28.1 Energy diagram summarizing the most used energy symbols and emergy indicators

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affected development patterns). From experience gained in previously investigated case studies, it appears that low ELRs (around 2 or less) are indicative of relatively low environmental impacts (or processes that are supported by large natural areas able to “dilute the impacts”). ELRs between 3 and 10 are indicative of moderate environmental impacts, driven by a country’s economic development (Brown and Ulgiati, 2001), while ELRs ranging from 10 up to extremely higher values indicate much heavier environmental impacts due to the larger flows of concentrated nonrenewable emergy in a relatively small local environment (i.e., without much possibility for dilution). If we combine the EYR (sensitive to the outside-vs.-local emergy-driven alternatives) and the ELR (sensitive to the equilibrium-vs.-disequilibrium emergydriven alternatives), we can generate an aggregated “sustainability” index (i.e., a measure of the potential exploitation of local resources—EYR—per “unit of loading” imposed on the local system—ELR). This indicator, named the Emergy Sustainability Index (ESI = EYR/ELR), is usefully applicable to quantify changes in openness and loading occurring over time in both technological processes and economies (Brown and Ulgiati, 1997). In principle, the lowest possible value of the ESI is 0 (when EYR = 0 and ELR ≠ 0 or when EYR ≠ 0 and ELR → ∞), while the theoretically upper limit (ESI → ∞) is only possible for untouched, mature ecosystems, if any, to which no investment F is provided. According to the results of several case studies investigated, ESI values lower than 1 are indicative of highly developed “consumer”-oriented countries, while higher ESI values are associated to still developing or poorly developed economies. The reason for such a surprising correlation (environmental sustainability poor economic development) must be found in the quantitative growth-oriented patterns, instead of qualitative-oriented patterns, that characterize mainstream economic thinking. The mainstream competitive economy is still, at present, the dominating strategy, as if policies were still unaware of the severe contradiction among available resources, increasing population, and increased aim toward improved lifestyles. As dictated by the Maximum Power Principle (Lotka, 1922a, 1922b; Odum, 1996)

[See entry 62], when resources are abundant, the maximum power output is ensured by competition for maximum resource use (no matter how efficient); while in the presence of a shrinking resource base, collaborative patterns that take the maximum advantage out of the decreased amount of emergy and share positive outcomes appear to be the most suitable sustainability strategy. Other performance indicators can be designed and applied to systems and processes, such as: ● Emergy Density, ED = U/Area (or the Empower Density, = U/Area/Time). This measures the total spatial and time concentration of resources and may suggest land to be a limiting factor, in other words, the need for supporting land around the system, for it to be sustainable. Higher ED values characterize city centers, information centers such as governmental buildings, universities and research institutions, power plants, and industrial clusters, while lower ED values are characteristic of rural areas and natural environments. ● Renewable Emergy Fraction (%Ren = R/U). The ratio of renewable emergy to total emergy use. In the long run, only processes with high %Ren are environmentally sustainable. ● Emergy per capita (Upc = U/population). This ratio of total emergy use to the total population can be used as a proxy for the standard of living of a region or a country.

Focus on economic systems and processes

Sustainability is a relative concept, showing several dimensions (carrying capacity, environmental, economic, social, cultural). Most often, a country’s development is driven by economic concerns, the easiest of which are addressed by policy makers and business operators. Therefore, worldwide and national economies tend to maximize GDP and welfare, with less or no focus on environmental challenges. “Absolute sustainability” concepts, with economies fully respecting biosphere constraints (resources, climate, biodiversity, etc.) can therefore be considered a reference target, to which “relative sustainability” performances are confronted. Therefore, a relative sustainability strategy Silvio Viglia and Sergio Ulgiati

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might develop processes and patterns which at least do not worsen (or slightly improve) the non-optimized emergy indicators (EYR, ELR, ESI, ED, etc.) of a country, toward more environmentally consistent production and consumption patterns (Brown and Ulgiati, 2001; Viglia et al., 2018). In so doing, at least conservative constraints could be suggested, as, for example: EYRprocess > EYRcountry; ELRprocess < ELRcountry; ESIprocess > ESIcountry In so doing, focus may be placed on land demand by a system or process, with more land needed as a source of renewable emergy (carrying capacity and environmental services for impacts dilution; Mellino and Ulgiati, 2015; Mellino et al., 2015). To monitor the relation of systems and processes on landscapes, Brown and Vivas (2005) designed a Landscape Development Intensity Index. Studies across various disciplines have employed the emergy analytical approach, such as regional and national sustainability (Ulgiati and Brown, 1998; Yang et al., 2010); natural ecosystems (Morandi et al., 2013; Odum and Odum, 2000); urban sustainability (Liu et al., 2009; Zhao et al., 2013); land use (Carey et al., 2011); trade (Huang et al., 2017; Rotolo et al., 2018). The latter topic (trade) is based on indicators of resource dependence Emergy Exchange Ratio (= Total imported emergy/Total exported emergy) and fair resource exchange, such as the Emergy Benefit Ratio (= Emergy of traded resource/ Emergy associated to money paid for), and raises issues of trade advantage, equitable development options, and shared responsibility of impacts. More recently, researchers paid attention to the integration of EMA with other assessment methods (life cycle assessment [LCA], ecological footprint, GIS, and economic analyses, among others; Mellino et al., 2014; Santagata et al., 2020).

Final remarks: the added value of the emergy approach

Compared to other environmental analysis methods, such as material flow analysis, the ecological footprint, and LCA, EMA maintains the advantage of unifying various dimensions, thereby linking systems of the natural environment and human economy Silvio Viglia and Sergio Ulgiati

(Amaral et al., 2016). Moreover, the novelty of the EMA is that most thermodynamic methods (except the exergy approach) do not recognize the difference in the quality of the various energy sources (Dong et al., 2008), so that one joule of heat is most often equated to one joule of fuel or electricity. Moreover, all the other methods in environmental assessment and ecological economics always estimate the value of ecosystem resources in terms that have been defined narrowly and anthropocentrically (user-side approach), while the EMA tries to capture the eco-centric value of ecological and manufactured products and processes (Hau and Bakshi, 2004), the so-called donor-side approach (i.e., focus is placed on the work of nature to provide resources). Nevertheless, like many groundbreaking ideas, the EMA has encountered much resistance and criticism, particularly from economists, physicists, and engineers since its inception in the 1980s. Some critics have focused on detailed practical aspects of the approach, while others have taken issue with specific parts of the theory and claims (Hau and Bakshi, 2004). For example, some economists have criticized the emergy analytical method from an anthropocentric perspective, claiming that unifying the values of various flows in a system disregards the fundamental principles of market demand and economic utility (Cleveland et al., 2000). Supporters of the emergy analytical method argue that the method is fundamentally opposed to the anthropocentric perspective and considers nature as a holistic system of which humans are only one of the components (Brown and Ulgiati, 1999, 2011). Efforts by many ecological-economists and others have been made to “internalize the externalities” (i.e., to adjust market valuation to give some consideration to ecosystem damages or use). According to Odum and Odum (2000, 21), “what is needed is the reverse: to ‘externalize the internalities’ to put the contributions of the economy on the same basis as the work of the environment” so that some economic processes should not be accepted and carried out if they are not within the dynamics and the limits of biosphere. As claimed by Huang and Odum (1991, 315), “money paid for resource inputs goes to the human extractors in large part for the work in obtaining those resources and not for the work of the environmental systems that

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produce them.” Instead, quantitative valuation based on the environmental contribution to human economy should be a prerequisite to any management decisions and to project evaluations concerning interactions between humans and nature; this is the main reason why EMA, and its metric, despite its practical limitations and the advancements that are still needed, provides a unique and complementary point of view in the wide ecological economics set of accounting methods. Silvio Viglia and Sergio Ulgiati

development intensity index. Environmental Monitoring and Assessment 101, 289–309. Campbell, E.T., Brown, M.T., 2012. Environmental accounting of natural capital and ecosystem services for the US National Forest System. Environment, Development and Sustainability 14, 691–724. Carey, R.O., Migliaccio, K.W., Li, Y.C., Schaffer, B., Kiker, G.A., Brown, M.T., 2011. Land use disturbance indicators and water quality variability in the Biscayne Bay Watershed, Florida. Ecological Indicators 11, 1093–1104. Cleveland, C., Kauffmann, R., Stern, D., 2000. Aggregation and the role of energy in the economy. Ecological Economics 32(2), 301–17. https://​doi​.org/​10​.1016/​S0921​-8009(99)00113​ References -5 Amaral, L.P., Martins, N., Gouveia, J.B., 2016. Dong, X., Ulgiati, S., Yan, M., Gao, W., 2008. A review of emergy theory, its application and Progress, influence and perspectives of emergy latest developments. Renewable and Sustainble theories in China, in support of environmentally Energy Reviews 54, 882–8. https://​doi​.org/​10​ sound economic development and equitable .1016/​j​.rser​.2015​.10​.048 trade. Energy Policy 36(3), 1019–28. https://​doi​ Brown, M.T., 2004. A picture is worth a thousand .org/​10​.1016/​j​.enpol​.2007​.11​.012 words: Energy systems language and simula- Hau, J.L., Bakshi, B.R., 2004. Promise and probtion. Ecological Modelling 178, 83–100. https://​ lems of emergy analysis. Ecological Modelling doi​.org/​10​.1016/​j​.ecolmodel​.2003​.12​.008 178(1–2), 215–25. https://​doi​.org/​10​.1016/​j​ Brown, M.T., Buranakarn, V., 2003. Emergy .ecolmodel​.2003​.12​.016 indices and ratios for sustainable material cycles Huang, S., An, H.H., Viglia, S., Buonocore, E., and recycle options. Resources, Conservation Fang, W., Ulgiati, S., 2017. Revisiting China– and Recycling 38, 1–22. Africa trade from an environmental perspective. Brown, M.T., Campbell, E., De Vilbiss, C., Journal of Cleaner Production 167, 553–70. Ulgiati, S., 2016. The geobiosphere emergy Huang, S., Odum, H.T., 1991. Ecology and baseline: A synthesis. Ecological Modelling economy: Emergy synthesis and public 339, 92–5. https://​doi​.org/​10​.1016/​j​.ecolmodel​ policy in Taiwan. Journal of Environmental .2016​.03​.018 Management 32(4), 313–33. https://​doi​.org/​10​ Brown, M.T., Cohen, M.J., Bardi, E., Ingwersen, .1016/​S0301​-4797(05)80069​-6 W.W., 2006. Species diversity in the Florida Lee, D.J., and Brown, M.T., 2021. Estimating the Everglades USA: A systems approach to calvalue of global ecosystem structure and proculating biodiversity. Aquatic Sciences 68(3), ductivity: a geographic information system and 254–77. emergy based approach. Ecological Modelling Brown, M.T., Ulgiati, S., 1997. Emergy-based 439, 109307. indices and ratios to evaluate sustainability: Liu, G.Y., Yang, Z.F., Chen, B., Ulgiati, S., Monitoring economies and technology toward 2009. Emergy-based urban health evaluation environmentally sound innovation. Ecological and development pattern analysis. Ecological Engineering 9(1–2), 51–69. https://​doi​.org/​10​ Modelling 220(18), 2291–2301. https://​doi​.org/​ .1016/​S0925​-8574(97)00033​-5 10​.1016/​j​.ecolmodel​.2009​.05​.019 Brown, M.T., Ulgiati, S., 1999. Emergy evaluation Lotka, A.J., 1922a. Contribution to the energetof the biosphere and natural capital. Ambio ics of evolution. Proceedings of the National 28(6), 486–93. Academies of Sciences 8(6), 147–51. Brown, M.T., Ulgiati, S., 2001. Emergy Measures Lotka, A.J., 1922b. Natural selection as a physof carrying capacity to evaluate economic ical principle. Proceedings of the National investments. Population and Environment: Academies of Sciences 8(6), 151–4. A Journal of Interdisciplinary Studies 22(5), Mellino S., Buonocore E., Ulgiati S., 2015. The 471–501. worth of land use: A GIS–emergy evaluation of Brown, M.T., Ulgiati, S., 2004. Emergy analysis natural and human-made capital. Science of the and environmental accounting. In: Encyclopedia Total Environment 506–07, 137–48. of Energy, C. Cleveland (ed), Academic Press, Mellino, S., Ripa, M., Zucaro, A., Ulgiati, S., Elsevier, Oxford, 329–54. 2014. An emergy–GIS approach to the evalBrown, M.T., Ulgiati, S., 2011. Understanding the uation of renewable resource flows: A case global economic crisis: A biophysical perspecstudy of Campania Region, Italy. Ecological tive. Ecological Modelling 223, 4–13. Modelling 271, 103–12. Brown, M.T., Vivas, M.B., 2005. Landscape

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168  Elgar encyclopedia of ecological economics Mellino, S., Ulgiati, S., 2015. Mapping the evolution of impervious surfaces to investigate landscape metabolism: an emergy–GIS monitoring application. Ecological Informatics 26(1), 50–59. Morandi, F., Campbell, D.E., Pulselli, R.M., Bastianoni, S., 2013. Using the language of sets to describe nested systems in emergy evaluations, Ecological Modelling, 265: 85–98. https://​doi​.org/​10​.1016/​j​.ecolmodel​.2013​.06​ .006 Odum, H.T., 1951. The stability of the world strontium cycle. Science 114(2964), 411–70. Odum, H.T., 1957. Trophic structure and productivity of Silver Springs Florida. Ecological Monographs 27, 55–112. https://​doi​.org/​10​ .2307/​1948571 Odum, H.T., 1971. Environment, Power and Society. John Wiley, New York. Odum, H.T., 1988. Self-organization, transformity, and information. Science 242, 1132–9. https://​doi​.org/​10​.1126/​SCIENCE​.242​.4882​ .1132 Odum, H.T., 1996. Environmental Accounting. Emergy and Environmental Decision Making. John Wiley & Sons, New York. https://​doi​.org/​ 10​.1017/​CBO9781107415324​.004 Odum, H.T., Odum, E., 2000. The energetic basis for valuation of ecosystem services. Ecosystems 3, 21–3. https://​doi​.org/​10​.1007/​ s100210000005

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Rotolo, G., Francis, C.A., Ulgiati, S., 2018. Environmentally sound resource valuation for a more sustainable international trade: Case of Argentine maize. Resources, Conservation & Recycling 131, 271–82. Santagata, R., Zucaro, A., Fiorentino, G., Lucagnano, E., Ulgiati S., 2020. Developing a procedure for the integration of life cycle assessment and emergy accounting approaches. The Amalfi paper case study. Ecological Indicators 117, 106676. Ulgiati, S., Brown, M.T., 1998. Monitoring patterns of sustainability in natural and man-made ecosystems. Ecological Modelling 108(1–3), 23–36. https://​doi​.org/​10​.1016/​S0304​ -3800(98)00016​-7 Viglia, S., Civitillo, D.F., Cacciapuoti, G., Ulgiati, S. (2018). Indicators of environmental loading and sustainability of urban systems. An emergy-based environmental footprint. Ecological Indicators 94, 82–99. Yang, Z.F., Jiang, M.M., Chen, B., Zhou, J.B., Chen, G.Q., Li, S.C., 2010. Solar emergy evaluation for Chinese economy. Energy Policy 38(2), 875–86. https://​doi​.org/​10​.1016/​j​.enpol​ .2009​.10​.038 Zhao, S., Song, K., Gui, F., Cai, H., Jin, W., Wu, C., 2013. The emergy ecological footprint for small fishfarm in China. Ecological Indicators 29, 62–7. https://​doi​.org/​10​.1016/​j​.ecolind​ .2012​.12​.009

29. Energy return on investment: a unifying principle for socio-ecological sustainability Energy return on investment (EROI), or occasionally energy return on energy/money/ water invested, is a metric that evaluates the efficiency of any given energy-gathering activity, usually in energy terms. EROI is a ratio of energy returned to society compared to the energy invested (or diverted) from society to get that energy. For example, a mean EROI of 20:1 for wind power means that you get ~20 units of electricity in return for every unit of energy invested in manufacturing, installing, maintaining, and decommissioning a wind energy system (Hall et al., 2014). Logically, EROI needs to be greater than one unit of energy returned from one unit of energy invested (the break-even point) for the activity to be viable. However, both organisms and societies need far more than to break even in real-world situations. Hall et al. (2009) estimated that a minimum EROI of 3:1 is required just to use oil in transport. If we consider the rest of society’s hierarchy of “energetic needs” (e.g., grow food, education, health, the arts, etc.), a much higher EROI is required (e.g., Lambert et al., 2014). Fundamentally, the concept of EROI provides a useful way of thinking about how organisms, ecosystems, and societies must obtain enough surplus energy returned from energy-gathering activities to live, reproduce, and thrive (Hall et al., 2009; Lambert et al., 2014; Hall, 2017). Energy surplus (or net energy) is the amount of energy left over after the costs of obtaining the energy have been accounted for. EROI has been used to develop the “net energy cliff” concept to show how EROI ratios below a 5:1 threshold impact society very strongly (Lambert et al., 2014). EROI is also an important concept to understand the different nature of energy flows from non-renewable stocks (e.g., coal, oil, natural gas, uranium, and thorium) and renewable funds (e.g., solar, wind, biomass, biofuels, hydroelectric, geothermal, and tidal), the importance of storage costs (e.g., Palmer, 2017; Kurland and Benson, 2019),

and other necessary indirect costs. Moreover, EROI can help us to overcome the false dualism between society and the rest of nature by informing how both require high-quality available energy to maintain their structures and functions, and it highlights the necessity to study socio-economic systems as embedded in the biophysical world of low-entropy energy and matter (Melgar-Melgar and Hall, 2020). There is no economy without available energy. Ecological and economic processes adhere to the first or conservation law of thermodynamics (which states that the total quantity of energy is conserved) and to the second or entropy law of thermodynamics (which states that the total quality of energy is not conserved; Melgar-Melgar and Hall, 2020). EROI is a useful metric for economics because it is based on these immutable physical laws rather than sometimes arbitrary or transitory human preferences. It is more powerful than conventional cost–benefit analysis for comprehensively understanding and comparing the energy trade-offs, efficiencies (and inefficiencies), resource depletion trends, resource quality, and surplus potentials of the different fuels and technologies that power, or might power, our socio-economic systems (Murphy et al., 2011; Melgar and Hall, 2020). Deriving EROI for carbonaceous fuels or other low-carbon energy technologies necessary to power society is simple in principle but may become complex in application. It is derived by dividing the energy returned to society over the energy invested by society to get that energy (Hall, 2011; Murphy et al., 2011; Hall and Klitgaard, 2018). The following equation represents its simplest form, although the details may vary according to the boundaries of the system in question: ​EROI ​=​

Energy returned from an energy gathering activity ___________________________________             ​ ​  Energy investment required to get that energy

or ​EROEI  ​=  _ ​ ER  ​ ​ EI

The basic steps to calculate an EROI are as follows: 1. State the objectives of your analysis to make sure that you have a clear understanding of the energy system you are

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trying to examine, and to learn and build from similar analyses already performed. 2. Create an Odum system flow diagram to help you understand the dynamics and boundaries of the energy system you are trying to analyze (e.g., Figure 29.1). 3. Establish the boundary of the energy system (fuel or technology) that you are planning to analyze. 4. Identify and gather the necessary data to estimate the energy returned (output) in the numerator. Usually the output is defined at point-of-extraction or the point-of-use. 5. Identify and gather the necessary data to estimate the energy invested (inputs) in the denominator. This should include direct (onsite) and indirect (energy used offsite but embodied in goods and services used onsite). 6. If energy data is unavailable, use monetary data and convert it to energy values using energy intensities of the country’s economy and sectors when available. 7. Calculate the EROI by dividing the energy retuned by the energy invested to obtain an EROI ratio. 8. Consider using a method for energy quality adjustment as part of sensitivity analysis that, at a minimum, corrects for the quality of electricity by multiplying it by a factor of 2.6 or 3 to represent mean thermal requirements. Include environmental and labor costs if appropriate and possible. For more detailed steps, see Murphy et al. (2011) and Hall (2017).

History of EROI: from biological to socio-economic applications

Fifty years ago, Hall (1972) explicitly coined the EROI concept while studying the energy cost and gains of migrating fish for his Ph.D. dissertation. He had been influenced by earlier socio-ecological analyses of net energy by Cottrell (1955), Boulding (1966), Lee (1969), Odum (1971), and others. EROI was originated as a means to describe the biological phenomenon of species needing to obtain enough energy surplus to be able to successfully survive and reproduce in dynamic ecosystems. Subsequently, amid the energy crises of the late 1970s and early 1980s, Hall and his colleagues formally applied the Rigo E.M. Melgar and Charles A.S. Hall

EROI term to fossil fuels (Hall et al., 1979; Hall et al., 1981; Hall and Cleveland, 1981), and to examine the biophysical foundation of the US economy (Cleveland et al., 1984; Hall et al. 1986). During the same period, Herendeen and Plant (1981) and Herendeen (1988) applied a version of EROI called “Energy Cost of Energy” to the analysis of renewable energy resources. In 1986, the evolution of the early applications of EROI, and its relationship to peak oil (Hubbert, 1956) and limits to growth (Meadows et al., 1972), became a pillar of the nascent field of biophysical economics, and subsequently ecological economics, culminating with the publication of Energy and Resource Quality: The Ecology of the Economic Process by Hall et al. (1986). A more updated version of this book by Hall and Klitgaard (2018), titled Energy and the Wealth of Nations: An Introduction to Biophysical Economics, continues to serve as a guide to understand the biophysical foundation of socio-economic systems (Melgar-Melgar and Hall, 2020). This biophysical foundation anchored in EROI is necessary to examine what could be the sustainable scale of energy sufficiency to power our socio-economic systems that our global ecosystems could sustain within planetary boundaries to enable the just distribution and efficient allocation goals of ecological economics (Steffen et al., 2015; Melgar and Burke, 2021). In the following decades, issues of energy scarcity were not as prevalent in the public eye as they had previously been during the energy shocks and subsequent inflationary crises of the 1970s–1980s, although there were warning signs of “The End of Cheap Oil” (Campbell and Laherrère, 1998). Since the early 2000s, the development of hydrofracking has enabled countries such as the US and Argentina to increase production of unconventional or tight oil. However, unconventional oil will also be depleted soon if the rate of drilling and extraction experienced in the past decade continues (Hughes, 2013; Heinberg, 2014). The financial crisis of 2007–08 (and the energy crisis from the Russian invasion of Ukraine in 2022) triggered renewed interest in biophysical economics and EROI analyses to explain the essential role of cheap energy in our economies, and how its scarcity can drive up inflation and lead to economic and financial

The figure is based on Hell et al. (1986) and Murphy et al. (2011). The legend of Odum’s Systems Language is based on Odum (1971, 1983, 2007).

Figure 29.1 An example of how to use Odum’s (1983, 2007) systems language symbols (including a legend of them that includes more symbols) to understand the energy system flows, boundaries, and biophysical economic interactions for an EROI analysis

Source:

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decline and even crises (e.g., Hall et al., 2008; Murphy and Hall, 2011; Tverberg, 2012). In 2011, Hall and Hansen (2011) edited a special issue of 21 articles on “New Studies in EROI” in the journal Sustainability, which initiated an explosion of new EROI publications that have aimed to standardize (e.g., Murphy et al., 2011), update (e.g., Hall et al., 2014), and expand applications of EROI (e.g., Court and Fizaine, 2017; Celi et al., 2018; Sgouridis et al., 2019; Capellán-Pérez et al., 2019). All of these newer studies using different techniques and data have come up with estimates for the EROI of modern societies and of its decline that are broadly similar to each other and to earlier values. EROI will continue to be relevant to guide the energy transition to low-carbon energy resources that humanity is attempting to embark on out of necessity to ameliorate climate change and the acceleration of fossil fuel depletion largely caused by our economic growth since the 1950s (Capellán-Pérez et al., 2014; Daly, 2014; Melgar-Melgar and Hall, 2020: Hagens, 2020; Laherrère et al., 2022).

A primer of EROI applications

The versatility of EROI as a net energy metric allows for flexible and varied applications based on the questions asked, boundaries chosen, and data availability for the analysis. While there have been criticisms that EROI boundaries and definitions have traditionally been too broad and variable to be useful for analytic work (e.g., King et al., 2015; Heinberg and Fridley, 2016), we believe that by paying careful attention to boundaries and definitions, and explicitly stating them, one can greatly reduce the differences in these literature values (e.g., Hall and Hansen, 2011; Hall et al., 2011; Hall, 2017). Most analyses have shown a trend of declining EROIs for oil and natural gas due to depletion, which means that many nations are nearing the so-called “net energy cliff” (Hall et al., 2014; Brockway et al., 2019). Although life cycle analysis (LCA) of energy is considered a parallel methodology to EROI, the former is used to quantify resource use/environmental impacts, while the latter is used today to measure the relationship of energy investments diverted from societal consumption to obtain energy for society. Murphy et al. (2016) call for the adoption of the LCA methodology to enhance rigor and standardiRigo E.M. Melgar and Charles A.S. Hall

zation in EROI analyses, while Arvesen and Hertwich (2015) call for caution in the use of LCA data to avoid the misclassification of energy flows. The following list of EROI applications shows how the concept has evolved from its early uses, and how it continues to be flexible enough for future improvements and applications according to different boundary settings and purposes. However, before we list the different applications of EROI, it is important to understand the three critical components of an EROI analysis, namely, the boundary, numerator, and denominator. Boundary The boundary of an EROI analysis should be chosen according to the purpose of the energy system analysis. Choosing the right boundary will also dictate data requirements and availability for the analysis. It is important to understand that, as you expand the boundary of your EROI analysis, the ratio will tend to decrease accordingly because you are incorporating more energy costs (inputs) or losses (subtracted from the output). But in general, the boundaries should be broad because it takes many activities to generate energy and much labor to do it. Prieto et al. (2012) suggest that the analysis “follow all money spent” under the assumption that all money spent requires energy to generate that good and service. Figure 29.2 shows a visual representation exemplifying the three basic boundaries of EROI analyses for oil. Numerator The numerator or energy output of an EROI analysis is usually straightforward and can be determined by multiplying the energy quantity produced by the energy content per unit. For alternative low-carbon renewable sources it can be done based on their generator nameplate capacity (energy generation under specific conditions as rated by the manufacturer) adjusted for intensity of use. Many costs should be related to normal economic pro-rating and depreciation analysis. Denominator The denominator or energy inputs of an EROI analysis can be more difficult to determine because it involves the consideration of a wide variety of direct and indirect energy

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Source:

Hall et al. (2014).

Figure 29.2 The three basic boundaries of EROI analyses and energy loss associated with the transformation of oil from extraction at the wellhead to consumer-ready fuel or gasoline

Source:

Hall and Klitgaard (2018).

Figure 29.3 The basic schematic model for energy used and gained for an energy system

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inputs/costs. The denominator is the most impacted by the boundary chosen in an EROI analysis, as we will discuss next. Figure 29.3 shows a schematic model of the boundary, numerator, and denominator in an EROI analysis.

Standard point-of-extraction EROI

A commonly used application is a standard point-of-extraction EROI, or EROIst/poe, which refers to energy directly won from nature (i.e., oil and natural gas at the wellhead, coal at the mine mouth, corn/sugarcane at the farmgate. The EROIpoe approach has the most basic boundary requirements, which include a simple standardized energy output (numerator) of a project divided by the sum of the direct (i.e., energy used on site) and indirect (i.e., energy used off site to make products used on site) energy inputs (denominator) used to generate the output. This basic standardized EROI approach is useful for comparing different fuels at the source of their generation. An EROIpoe is recommended as a starting point for analysts to avoid issues that may arise from choosing different boundaries as you move along the energy food chain of society. The following is the basic equation for a standard EROI: ​EROI​poe   ​ = ​ ​

Energy returned to society at point of extraction

__________________________________       ​ ​  Direct  &  indirect energy invested to get energy

The following applications for point of use, extended, and societal EROI have been derived by expanding the system’s boundary, which tends to decrease the energy delivered to society as energy costs are added to the denominator.

Point-of-use EROI

A point-of-use EROI, or EROIpou, can be used to understand how much energy is needed to deliver energy to society to get the job done (e.g., a gas station, charging station). An EROIpou is the next step after an EROIpoe in the energy food chain of society because it extends the boundary to include the energy costs required to get plus deliver (e.g., refining and transporting oil) energy to the point-of-use for socio-economic activities. Murphy et al. (2022) have recently argued that using point-of-use EROIs rather than Rigo E.M. Melgar and Charles A.S. Hall

point-of-extraction EROIs avoids inaccurate comparisons between thermal fuels and electricity-producing technologies because the energy costs of thermal fuels increase from poe to pou lowering their EROI. The following is the basic equation for a point-of-use EROI: ​EROI​pou   ​ ​=​

Energy returned to society at point of use _______________________________________            ​ ​  Direct & indirect energy invested to get & deliver energy

Extended EROI Since societies care about the energy services (e.g., drive a truck, power a computer) and not about energy, per se, an extended EROI, or EROIext, is recommended to correct an EROIpou for the energy costs required to create and maintain the infrastructure to get and deliver energy at the point of use. An EROIext comprehensively assess how much upstream (i.e., energy costs of finding and producing energy) and downstream (i.e., energy required to deliver the energy service) energy we need to invest to use and consume energy in socio-economic activities to generate well-being. An EROIext is the third step after EROIpoe and EROIpou in the energy food chain of society because it extends the boundary to include the energy costs required to get plus deliver plus use energy (e.g., Figure 29.2). The following is the basic equation for an extended EROI: ​EROI​ext   ​ ​=​ Energy returned to society as services

            ​ ​ _____________________________________ Direct & indirect energy invested to get, deliver, & use energy

Societal EROI

Ultimately, a societal EROI, or EROIsoc, is recommended to understand the macroeconomic significance of EROI for a nation or society by estimating all the energy gains and costs of obtaining energy. The following is the basic equation for a societal EROI: ​EROI​soc   ​ ​=​ Summation of energy content of all energy delivered to society __________________________________              ​ ​  Summation of all energy invested to get energy resources

In these analyses we are still being conservative in estimating costs: Should we include

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the energy costs of supporting the laborer’s paycheck? The costs of educating and training engineers? Of maintaining roads used to deliver steel or remove oil? All of these are necessary costs that are never included, so all estimates are probably too high! Future EROI analyses should attempt to incorporate these energy costs.

Relating financial and energy ROI

King and Hall (2011) developed a theoretical framework based on empirical information for relating EROI with monetary return on investment to understand the underlying physical relationship between price and EROI, and how long-term declining EROI trends might impact energy-producing entities. They found a fairly strong negative correlation between EROI and prices for oil and for natural gas. Wang et al. (2019) have created a model showing that combining financial and energy ratios provides more comprehensive and accurate information than either one alone in their analysis of the economic and biophysical sustainability of oil sands, shale oil, and shale gas operations in North America. Moreover, Wang et al. (2019) recommend that the EROI indicator and their comprehensive efficiency indicator should be included and audited in the sustainability disclosure reports of companies such as the Global Reporting Initiative to provide incentives for companies to innovate, improve efficiency, and meet the energy and environmental objectives of public policy. Recently, Oosterom and Hall (2022) applied the EROI concept to supplant or supplement traditional discounted cash-flow analysis (DCF) to enhance the evaluation of energy investment decisions when considering the increasingly popular environmental, social, and governance (ESG) investment criteria. They demonstrate that EROI can provide a more robust understanding of energy investments than DCF alone, using the examples of oil sands vs. microbial-enhanced oil recovery.

Other creative applications of EROI

Recently, other creative applications of EROI have been developed to address issues of long-term dynamics of net energy (e.g., Dale et al., 2012; Court and Fizaine,

2017; Capellán-Pérez et al., 2019), water requirements of energy production though estimates of energy return on water invested (Mulder et al., 2010), opportunities for efficiency improvements of low-carbon renewable technologies (Atlason and Unnthorsson, 2014), EROEI considerations in input–output macroeconomic modeling and its implications for economic growth under an energy transition scenario (Fagnart and Germain, 2016), importance of connecting EROI with the price of energy and other goods and services (Herendeen, 2015), EROI comparisons of renewables and carbon capture and sequestration (Sgouridis et al., 2019; Sekera and Lichtenberger, 2020), examination of the limits of renewable energy production (Dupont et al., 2018, 2020), the combination of EROI with the ecological footprint of renewables (Ward et al., 2020), EROI of global agriculture, aquaculture, fishing, and forestry energy systems (Marshall and Brockway, 2020), and analogous application measures for energy stored and energy saved on investment. Next, we explore some of these applications in more detail with an emphasis on the energy transition. Dynamic EROI A dynamic function of EROI, or EROIdyna, has been proposed as a “top-down” approach to analyze and predict the long-term dynamics of energy resources in relation to technological improvements and depletion of fossil fuels (Dale et al., 2011). An EROIdyna is considered necessary because most EROI estimates tend to be static by accounting for a particular location and time through a “bottom-up” approach. Dale et al. (2012) incorporated an EROIdyna function in the development of the global energy model using a biophysical approach (GEMBA) to understand the dynamics of EROI in relation to how it increases in early stages of investments and technological improvements, and then how it tends to peak and decline due to diminishing returns. Court and Fizaine (2017) proposed and applied a new theoretical dynamic expression of EROI based on Dale et al. (2011) and using a price-based approach based on King and Hall (2011) in their global analysis of past and future maximum EROIs for oil, natural gas, and coal. Their analysis shows that maximum EROIs for oil and natural gas occurred in the Rigo E.M. Melgar and Charles A.S. Hall

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1930s–40s, and that coal’s maximum EROI could peak between 2025–45. Recently, Capellán-Pérez et al. (2019) have developed their own comprehensive EROIdyna methodology for the MEDEAS-World model to examine the EROI and material implications of transitioning the global electricity sector from fossil fuels to renewable technologies. The MEDEAS-W model shows that the large, up-front energy investment requirements of such a transition would substantially decrease the EROIs of low-carbon technologies by mid-century and lead to a re-materialization of the economy, putting into question the blind pursuit of “green” growth. Ideal EROI Atlason and Unnthorsson (2014) proposed using an ideal EROI, or EROIide, to estimate the potential of EROI improvement for renewable low-carbon energy technologies such as hydropower, geothermal, and wind. The authors acknowledge that, due to upper limits in efficiency, an EROI of an energy system can never be equal to an EROIide. However, they believe that their EROI application can be useful for energy policy makers and planners by providing an understanding of the difference between an EROIstnd and an EROIide. This difference needs to be large enough to invest in research and development to achieve EROI improvements, which will become essential as nations embark on decarbonization. The following is the basic equation for an ideal EROI: ​EROI​ide   ​ = ​ ​

Theoretical maximum output omitting all losses __________________________________            ​ ​  Direct & indirect energy investment to get energy

Relating EROI and ecological footprint

Ward et al. (2020) combined EROI with the ecological footprint concept to develop the renewable energy equivalent footprint (REEF) to analyze the biocapacity implications of meeting humanity’s energy demands with renewable energy. The authors apply their REEF framework to biomass fuel production and solar electricity production, and use EROI to model the energy investments needed in cross-subsidization between the two renewable energy sources to quantify Rigo E.M. Melgar and Charles A.S. Hall

the footprint of their supply. REEF is presented as a complementary tool to the more conventional carbon footprint and can be applied at different scales, including for an energy-consuming activity, community, country, or the world.

EROI and energy transitions

As countries embark on an energy transition to low-carbon energy, it will be vital to continue to understand what is the minimum/ sufficient EROI threshold for society (Hall et al., 2009; Lambert et al., 2014; Fizaine and Court, 2016), and the biophysical and societal implications of the large, up-front energy investments and material requirements that will be needed to achieve it (Sers and Victor, 2018; King and van den Bergh, 2018; Capellán-Pérez et al., 2019; de Castro et al., 2020). Recently, Brockway et al. (2019) estimated the global final-stage (e.g., electricity, gasoline, gas) EROI for fossil fuels to demonstrate that a final EROI moves closer to the EROI of renewable energies (as opposed to a primary-stage EROI), suggesting that an energy transition might be biophysically possible, and that depletion is decreasing the EROI of fossil fuels. But Capellán-Pérez et al. (2019) and Dupont et al. (2021) found in their models that transitioning to 50 percent or 100 percent penetration of renewables might lead to huge capital requirements and very low EROIs. Another aspect of an energy transition is the role of technologies such as carbon capture and storage (CCS) in enabling society to reduce emissions while continuing to burn carbon. However, Sgouridis et al. (2019) demonstrated that investing in renewables plus storage has a higher EROI than constructing CCS fossil fuel power stations. Ecological economists are increasingly studying the energy transition risks on society and the economy by incorporating EROI in their ecological macroeconomic models. For example, the model by Jackson and Jackson (2021) showed that the potential negative macroeconomic effects of the transition, namely recession, stagnation, stagflation, increasing inequality and asset stranding, are positively related to the capital intensity of low-carbon energy and declining EROI. Furthermore, Jacques et al. (2023) recently developed a biophysical stock-flow consistent model that incorporates EROI to assess

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the economic consequences of an energy transition to low-carbon energy, showing that such a transition might be easier in a post-growth economy scenario. These examples of EROI applications show how important EROI is to examine and understand the implications of energy transitions for the well-being of society and the rest of nature amid limits to growth (Hall, 2022). From an ecological economics perspective, an energy transition needs to avoid feeding into the surplus-seeking, self-reinforcing and expansionary “economic superorganism” (Gowdy, 2021; King, 2020; Krall, 2022) by enabling the degrowth of wasteful and inefficient energy-materials throughput in the economy with policies informed by EROI that prioritize energy sufficiency within planetary boundaries (Burke and Melgar, 2022). In summary, we argue that EROI is a critically useful concept to help understand the patterns of, and interconnections among, the environment, energy, and socio-economic development, and to inform future energy transitions. Other EROI resources can be found at: the Journal of Biophysical Economics and Sustainability, the BioPhysical Economics Institute, and the International Society for BioPhysical Economics. Diagrams.net is an open-access modeling tool to create Odum’s systems diagrams for EROI analyses or other systems modeling of energy flows. Rigo E.M. Melgar and Charles A.S. Hall

References

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30. Energy transition(s) Introduction

Social scientists, among others, have appended the concept of “transition” to processes or periods of changing from one state or condition to another (e.g., demographic transition, political transition, transition economies, ecological transition). When applied to energy, it has, in principle, the same implication: any meaningful change from one state of an energy system to another may constitute an energy transition. These alterations might affect the amounts of energy, the quality (energetic, economic, environmental, etc.), the forms of energy harnessed, how these are converted (or not) and delivered to the final consumers, as well as the changes in the quantity, quality, and variety of the energy services provided. These changes are intertwined, although some are easier to identify and measure than others. Precisely because these processes are unlikely to happen one at a time, and because there may be variations across time and space, scholars prefer the use of the concept in the plural form: energy transitions. Historians have long studied energy transitions as part of their scrutiny of the past. This was the first evocation of the concept. A different perspective, looking into the future of the energy systems, arose first in the 1970s— due to the oil crisis—and later from the 1990s, with increasing concern over global climate impacts of energy-related human activities. These led academics and public policy makers to increase their interest in accelerating the changes of the energy systems toward more desirable scenarios (away from petroleum in the 1970s, to a low-carbon economy from the 1990s) by managing the current and future energy transitions. This entry differentiates the positive from the normative approaches to the concept of energy transition(s). Historical descriptions and analysis of past energy transitions study how, when, and where energy systems transformed in the past but, for the most part, contain no indication of approval or disapproval. In contrast, when looking into the future of energy systems, the energy transition has come to have a precise meaning of what is the desirable objective—a fast pathway to a low-carbon energy system—and

focuses on how best achieve the set targets, thus transforming the concept into a normative one. The two meanings (positive and normative) of energy transition(s) are present in the literature, but the normative form is by far the most widely used by non-academics; within academia, most of the literature on energy transitions now also has a normative focus. Research about the past and future of energy systems conforms to two distinct but interacting bodies of knowledge about the energy transition(s).

Energy transitions in the past

American philosopher and historian of technology Lewis Mumford published a book in 1934 that reviewed history from an energetic viewpoint for the first time; his work followed the ideas of Geddes, who possibly was the first to interpret history as an understanding of the human in its environment, whether natural world or urban spaces, turning the physical world into an important part of the explanation of how humans behave and relate (Martínez-Alier & Schlüpmann, 1991). In Munford’s view, history could be interpreted in terms of successive episodes of “energy releases.” Each of them would provide more energy for society, an improvement in the supply regularity, more flexibility in the distribution, and more efficient use. Similarly, Cottrell (1955), a sociologist, described the evolution of social and economic change in terms of energy, emphasizing the importance of energy transitions, as the shift from animate energy sources (human labor and draft animals) to inanimate energy sources and their associated converters (fossil fuels, steam, and the internal combustion engine). Economic historians such as Cipolla (1962) and Wrigley (1962) would reformulate some of these ideas. Both scholars emphasized it was essential for society to break free from the constraint imposed by the energy budgets of organic economies, which depended almost exclusively on annexing as much as possible of the annual inflow of solar energy from plants, humans, and animals. The historical significance of these changes, especially from the development of the steam engine, is that humanity progressively obtained higher levels of disposable energy per head. Part of this translated only into more energy consumption per capita for different uses (heating, lighting, transports,

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etc.) but also into more energy per worker, and consequently, greater labor productivity (Cipolla, 1962). As a result of these, industrial societies entered a new cycle of economic growth, which at the same time acted as a stimulus for the development of new energy forms: “the more energy produced, the more energy was sought out” (Cipolla, 1962, 63). However meaningful, the work of historians did not provide quantitative precision or a specific chronology to their observations about the alterations of the energy systems in the very long run. The primary reason for this was that energy transition quantification required long-term energy data (over a century or more). The earlier quantitative reconstructions of historical energy data sets described the energy paths followed by affluent nations.1 From these opulent countries’ historical energy data was derived a set of common features regarding pace, irreversibility, and sequence within energy transitions. Altering the energy basket revealed a slow process. In several of the crucial stages, such as the passage from organic to mineral energy or the switch from coal to oil, the duration of the process ranged from several decades to over a century. The available evidence also indicated that past transitions followed

an irreversible progression. The energy ladder hypothesis, by which countries adopt higher-quality energy sources—cleaner and more efficient—as their income increases, seems to imply a path toward increasing energy mix diversification over time as countries become richer (Rubio-Varas & Muñoz-Delgado, 2019), given that the lower-quality energy sources tend to remain in the energy basket. In the Western world, there was also an apparent sequence among the countries that first reached the higher rungs on the ladder: wealthier economies typically reached upper levels of the energy ladder ahead of poorer economies (Figure 30.1). Consequently, future transitions were expected to be slow, with today’s most affluent economies setting the pace while the rest of the world climbed the energy ladder, step by step, behind the leaders (Fouquet, 2010; Smil, 2021; Murphy, 2001; Gibbons et al., 1991; Lönnroth et al., 1980). As more data became available for other parts of the world, alternative historical types of energy transitions for poorer and rather small energy consuming countries emerged that did not fit the standards of the historical experience of the rich nations. Henriques (2011) showed that Portugal leapfrogged from firewood as the predominant fuel, straight into oil by the mid-1960s (Figure

Figure 30.1 The primary energy transition in England and Wales, 1800–2008

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Energy transition(s)  183

30.2). Rubio and Folchi (2012) reconstructed the data for 20 Latin American countries for over a century and showed that small energy consumers had earlier (mid-1930s) and faster (less than a decade) transitions from coal to oil than the Western countries did. Henriques and Sharp (2015) found a quick transition from firewood to coal in Denmark, another small energy consuming country. These results fit better with the evidence that during the past 30 years, developing countries made energy transitions earlier, faster, and with greater diversity than had been previously understood, although they still lagged behind the wealthy economies (Marcotullio & Schulz, 2007). This transition “acceleration” from early to late adopters has also been documented for transport systems (Grubler, 1996). Early adopters face formidable challenges in transitioning to newer transport systems due to high sunk costs. Among the factors explaining slow transitions are scale, or market size (implying that larger systems change more slowly than smaller ones), and technological interrelatedness and infrastructure needs (i.e., the

more infrastructure-intensive and complex technology systems are, the slower they can be changed; Grubler, 2012). Since the quantity and the structure of the energy supply were quantifiable variables, the energy transition has been too often simplified to the moment when a predominant source of energy reduces its share in the energy mix and another takes its place as the principal one. Even when the energy mix crucially affects energy efficiency, energy intensity, energy security, and carbon intensity of a country (Kander et al., 2020), the amount of energy at our disposal and the structure of the energy mix are just two of the multiple characteristics defining the state of energy systems. Energy systems have, over time, endured enormous transformations in the quality, the methods of conversion and delivery, and the destination of final energy. Yet, we have less quantitative evidence about the evolution of these other alterations. A limitation has arisen, mostly from the much sparser historical records on energy end-use (Grubler, 2012), with the notable exception of Fouquet (2008, 2010), who studied the

Figure 30.2 The primary energy transition in Portugal, 1856–2008

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long-term evolution of energy services in the UK for light, mobility, power, and thermal energy end-uses integrating energy accounts with efficiency and costs of energy service provision. Although there exist some global generalizations of the evolution of final energy consumption (Smil, 2021; Grubler, 1998), these lack the finesse and precision of historical case studies. Consequently, the prerogative of energy demand over energy supply as driver of transitions remains elusive to quantify in historical terms.

Energy transitions in the future

The use of the term “energy transition” in public discourse and policy was popularized by US President Jimmy Carter in a televised speech from the Oval Office in 1977; in it, he called to “look back into history to understand our energy problem”—at the time, the rise of oil prices. The consequences of the oil crisis for the future of energy systems and how to accelerate the move away from oil became both an item on the political agenda and an object of academic research (Attiga, 1978; Clapp, 1980; Lönnroth et al., 1980; Gheorghe, 1981). As is often the case in science, theory and models follow empirical understanding and data rather than the other way around: the shape and pace of future transitions was first interrogated by looking at transitions of the past (Arapostahis & Pearson, 2019; Fouquet & Pearson, 2012; Grubler, 2012; Bennett, 2012; Pearson & Foxon, 2012; Allen, 2012; Steinmueller, 2013; Napp et al., 2017). Toward the end of the 20th century, the need to address climate change became the major driver for a transition from an energy sector dominated by fossil fuels to one based on renewable energy sources. The energy sector is currently the main emitter of greenhouse gases. Thus, transforming the energy system in a very specific direction—toward low-carbon technologies—has become the objective of a wide range of researchers, and its own literature branch has developed: transition studies. The energy system has become one of the core examples of the socio-technical transitions (STT) literature, as a system made up of a wide range of analytically separable but dynamically inter-related areas—infrastructures, technology, corporate groups, user practices, civil society, institutions, politics, Mar Rubio-Varas

and ultimately the environment (Foxon, 2011; Rotmans et al., 2001). According to Markard et al. (2012), four frameworks achieved some prominence in transition studies. These include transition management (Kern & Smith, 2008; Loorbach, 2010; Rotmans et al., 2001), strategic niche management (Kemp et al., 1998; Raven & Geels, 2010; Smith, 2007), the multilevel perspective on STSs (Geels, 2002; Geels & Schot, 2007; Smith et al., 2010), and technological innovation systems (Bergek et al., 2008; Jacobsson & Johnson, 2000; Hekkert et al., 2007). The conceptual frameworks of STT are often found to be difficult to operationalize in quantitative energy analyses that meet policy development requirements. The quantitative modeling of socio-technical energy transitions (STET), which merges the conceptual frameworks of STT with energy modeling attempts to overcome these hurdles (Li et al., 2015). In part because of the above literature and because of the policy interest about the future of energy systems, the term “energy transition” has come to have a specific meaning referring to the global energy sector’s shift from fossil-based systems of energy production and consumption—including oil, natural gas, and coal—to an energy system dominated by non-emitting technologies (often identified as renewables, leaving out nuclear on most occasions). For the public, the press, and most stakeholders, the switch to a renewable energy system has become “the energy transition.” Defining who should be the agents driving this future energy transition (the market, the government, and/or the citizens)—that is, the governance of the desirable energy transition—has created a wide scholarly debate and policy developments including calls for “energy democracy” and active forms of “energy citizenship.” While energy citizenship tends to emphasize behavior change and ways for individuals to participate in energy systems, energy democracy tends to focus on institutionalization of new forms of participative governance, and often places collectives as central agents of change (Wahlund & Palm, 2022). Because the future transition out of fossil fuels is certainly not a guarantee for a fairer world (Welton & Eisen, 2019; Sovacool et al., 2019), the justice of the energy transition has also become a central piece of the aca-

Energy transition(s)  185

demic and policy discussion. But what will constitute a “just energy transition” is also a multifaced question (Wang & Lo, 2021). Adding to the difficulties, a large part of the ecological perspective on the energy transition focuses on the biophysical constraints (IEA, 2021), questioning the feasibility of the future energy transition (de Blas et al., 2021) within planetary boundaries (Rockström et al., 2009). Mar Rubio-Varas

Note 1.

The following countries were examined: Austria (Krausmann & Haberl, 2002), Canada (Steward, 1978; Unger & Thistle, 2013), Italy (Malanima, 2006), Germany and France (Kander et al., 2014), Netherlands (Gales et al., 2007), Portugal (Henriques, 2011), Spain (Rubio, 2005), Sweden (Kander, 2002), United Kingdom (Fouquet & Pearson, 1998; Fouquet, 2010; Warde, 2007), United States (Schurr & Netschert, 1960; O’Connor & Cleveland, 2014).

References

Allen, R.C. (2012). Backward into the future: the shift to coal and implications for the next energy transition. Energy Policy 50, 17–23. Attiga, A. (1978). The impact of energy transition on the oil-exporting countries. Journal of Energy and Development 4(1), 41–8. Arapostathis, S.G., & Pearson, P.J. (2019). How history matters for the governance of sociotechnical transitions: An introduction to the special issue. Environmental Innovation and Societal Transitions 32, 1–6. Bennett, S.J. (2012). Using past transitions to inform scenarios for the future of renewable raw materials in the UK. Energy Policy 50, 95–108. Bergek, A., Jacobsson, S., Carlsson, B., Lindmark, S., & Rickne, A. (2008). Analyzing the functional dynamics of technological innovation systems: A scheme of analysis. Research Policy 37, 407–29. Cipolla, C.M. (1962). The Economic History of World Population. Harmondsworth: Penguin Books. Clapp, N. (1980). Energy transition in the United States – Low-head resource and its economic potential. Natural Resources Forum 4(1), 109–14. Cottrell, W.F. (1955). Energy and Society: The Relation Between Energy, Social Change, and Economic Development. New York: McGraw Hill. de Blas, I., Mediavilla, M., Capellán-Pérez, I., & Duce, C (2021), The limits of transport decarbonization under the current growth paradigm.

Energy Strategy Reviews 32, 100543. https://​ doi​.org/​10​.1016/​j​.esr​.2020​.100543 Foxon, T.J. (2011). A coevolutionary framework for analysing a transition to a sustainable low carbon economy. Ecological Economics 70, 2258–67. Fouquet, R. (2008). Heat, Power and Light: Revolutions in Energy Services. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Fouquet, R. (2010). The slow search for solutions: Lessons from historical energy transitions by sector and service. Energy Policy 38(11), 6586–96. Fouquet, R., & Pearson, P.J.G. (1998). A thousand years of energy use in the United Kingdom. Energy Journal 19(4), 1–41. Fouquet, R., & Pearson, P.J.G. (2012). Past and prospective energy transitions: Insights from history. Energy Policy 50, 1–7. Gales, B., Kander, A., Malanima, P., & Rubio, M.D.M. (2007). North versus South: Energy transition and energy intensity in Europe over 200 years. European Review of Economic History 11(2), 219–53. Geels, F.W. (2002), Technological transitions as evolutionary reconfiguration processes: A multi-level perspective and a case-study. Research Policy, 31, 1257–74. Geels, F.W., & Schot, J. (2007). Typology of sociotechnical transition pathways. Research Policy, 36(2007), 399–417. Gheorghe, A. (1981). Energy transition policies to meet the needs of a medium developed country (Romania). Materials and Society 5(1), 89–113. Gibbons, J., Blair, H., & Peter, D. (1991). US energy transition: On getting from here to there. Physics Today, 44(7), 22–30. Grubler, A. (1996). Time for a change: On the patterns of diffusion of innovation. Daedalus 125(3), 19–42. Grubler, A. (1998). Technology and Global Change. Cambridge: Cambridge University Press. Grubler, A. (2012). Energy transitions research: Insights and cautionary tales. Energy Policy 50, 8–16. Hekkert, M., Suurs, R.A.A., Negro, S., Kuhlmann, S., & Smits, R. (2007). Functions of innovation systems: A new approach for analysing technological change. Technological Forecasting and Social Change, 74 (2007), 413–32. Henriques, S.T. (2011). Energy transitions, economic growth and structural change: Portugal in a long-run comparative perspective. Thesis published in Lund Studies in Economic History, vol. 54. Lund University. Henriques, S.T., & Sharp, P. (2015). The Danish agricultural revolution in an energy perspective: A case of development with few domes-

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186  Elgar encyclopedia of ecological economics tic energy sources. Economic History Review 69(3), 844–69. International Energy Agency (2021). The Role of Critical World Energy Outlook Special Report. IEA. Jacobsson, S., & Johnson, A. (2000). The diffusion of renewable energy technology: An analytical framework and key issues for research. Energy Policy 28, 625–40. Kander, A. (2002). Economic growth, energy consumption and CO2 emissions in Sweden 1800–2000. Thesis published in Lund Studies in Economic History, vol. 19, Lund University. Kander, A., Malanima, P., & Warde, P. (2014). Power to the People: Energy in Europe over the Last Five Centuries. Princeton, NJ: Princeton University Press. Kander, A., Stern, D.I., & Rubio-Varas, M. (2020). Energy intensity: The roles of rebound, capital stocks, and trade, in M. Ruth (ed), A Research Agenda for Environmental Economics, 122–42. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Kemp, R., Schot, J., & Hoogma, R. (1998). Regime shifts to sustainability through processes of niche formation: The approach of strategic niche management. Technology Analysis & Strategic Management 10, 175–95. Kern, F., & Smith, A. (2008). Restructuring energy systems for sustainability? Energy transition policy in the Netherlands. Energy Policy 36(2008), 4093–4103. Krausmann, F., & Haberl, H. (2002). The process of industrialization from the perspective of energetic metabolism. Socioeconomic energy flows in Austria 1830–1995. Ecological Economics 41(2), 177–201. Li, F.G.N., Trutnevyteb, E., & Strachana, N. (2015). A review of socio-technical energy transition (STET) models. Technological Forecasting and Social Change 100, 290–305. Lönnroth, M, Steen, P., & Johansson Thomas, B. (1980). Energy in Transition: A Report on Energy Policy and Future Options. Berkeley: University of California Press. Loorbach, D. (2010). Transition management for sustainable development: A prescriptive, complexity-based governance framework. Governance 23, 161–83. Markard, J., Raven, R., & Truffer, B. (2012). Sustainability transitions: An emerging field of research and its prospects. Research Policy 41(6), 955–67. Malanima, P. (2006). Energy Consumption in Italy in the 19th and 20th Centuries: A Statistical Outline. Naples: ISSM-CNR. Marcotullio, P.J., & Schulz, N.B. (2007). Comparison of energy transitions in the United

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States and developing and industrializing economies. World Development 35(10), 1650–83. Martínez-Alier, J., & Schlüpmann, K. (1991). La Ecología y la Economía. México D.F.: Fondo de Cultura Económica. Mumford, L. (1934). Technics and Civilisation. New York: Harcourt, Brace and Company Murphy, J.T. (2001). Making the energy transition in rural east Africa: Is leapfrogging an alternative? Technological Forecasting and Social Change 68(2), 173–93. Napp, T., Bernie, D., Thomas, R., Lowe, J., Hawkes, A., & Gambhir, A. (2017). Exploring the feasibility of low-carbon scenarios using historical energy transitions analysis. Energies 10(1), 116. O’Connor, P.A., & Cleveland, C.J. (2014). U.S. energy transitions 1780–2010. Energies 7(12), 7955–93. Pearson, P.J.G., & Foxon, T.J. (2012). A low carbon industrial revolution? Insights and challenges from past technological and economic transformations. Energy Policy 50, 117–27. Raven, R., & Geels, F.W. (2010). Socio-cognitive evolution in niche development: Comparative analysis of biogas development in Denmark and the Netherlands (1973–2004). Technovation 30, 87–99. Rockström, J., Steffen, W., Noone, K., Persson, Å., Chapin, III, F.S., Lambin, E., Lenton, T.M., et al. 2009. Planetary boundaries: Exploring the safe operating space for humanity. www​ Ecology and Society 14(2), 32. http://​ .ecologyandsociety​.org/​vol14/​iss2/​art32/​ Rotmans, J., Kemp, R., & van Asselt, M. (2001). More evolution than revolution: Transition management in public policy. Foresight 3(1), 15–31. Rubio, M.D.M. (2005). Energía, economía y CO2: España 1850–2000. Cuadernos Económicos de ICE 2(70), 51–71. Rubio, M.D.M., & Folchi, M. (2012). Will small energy consumers be faster in transition? Evidence from the early shift from coal to oil in Latin America. Energy Policy 5, 50–61. Rubio-Varas, M., & Muñoz-Delgado, B. (2019), Long-term diversification paths and energy transitions in Europe. Ecological Economics 163, 158–68. Schurr, S.H., & Netschert, B.C. (1960). Energy in the American Economy, 1850–1975. An Economic Study of its History and Prospects. Baltimore: Johns Hopkins University Press. Smil, V. (2021). Grand Transitions: How the Modern World Was Made. New York: Oxford University Press. Smith, A. (2007). Translating sustainabilities between green niches and socio-technical regimes. Technology Analysis & Strategic Management 19, 427–50. Smith, A., Voß, J.-P., & Grin, J. (2010). Innovation studies and sustainability transitions: The allure

Energy transition(s)  187 of the multi-level perspective and its challenges. Research Policy 39, 435–48. Sovacool, B.K., Martiskainen, M., Hook, A., & Bake, L. (2019). Decarbonization and its discontents: A critical energy justice perspective on four low-carbon transitions. Climate Change 155, 581–619. Steinmueller, W.E. (2013). The pre-industrial energy crisis and resource scarcity as a source of transition. Research Policy 42(10), 1739–48. Steward, F.R. (1978). Energy consumption in Canada since confederation. Energy Policy 6(3), 239–45. Unger, R.W., & Thistle, J. (2013). Energy Consumption in Canada in the 19th and 20th Centuries. A Statistical Outline. Napoli: ISSM-CNR. Wahlund, M., & Palm, P. (2022). The role of energy democracy and energy citizenship for

participatory energy transitions: A comprehensive review. Energy Research & Social Science 87, 102482. https://​doi​.org/​10​.1016/​j​.erss​.2021​ .102482 Wang, X., & Lo, X. (2021). Just transition: A conceptual review. Energy Research & Social Science 82, 102291. https://​doi​.org/​10​ .1016/​j​.erss​.2021​.102291 Warde, P. (2007). Energy Consumption in England and Wales 1560–2000. Naples: ISSM-CNR. Welton, S., & Eisen, J. (2019), Clean energy justice: Charting an emerging agenda. Harvard Environmental Law Review 43,307. Wrigley, E. A. (1962). The supply of raw materials in the Industrial Revolution. Economic History Review XV, 1–16.

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31. Entropy The First Law of Thermodynamics tells us that energy can be neither created nor destroyed. The Second Law places additional limits on energy transformations and reflects qualitative characteristics. It states that energy can only be transformed by the consumption of quality. Locally, the quality can be improved, but this can only occur at the expense of a greater quality deterioration elsewhere. The level of quality deterioration, frequently associated with disorder, is measured through the property “entropy”, denoted with S. Entropy, like energy, is a state function and can be obtained by knowing the intensive properties of a system (i.e. pressure, temperature, and composition). Mathematically, the change of entropy ​∆ S​of an internal reversible process (simply put, ideal process, where no irreversibilities occur within the system) is expressed as:

entropy generation. In other words, and as stated before, if ∆ ​  S ​< 0, it means that, despite the irreversibilities created in the process, the quality of the system has been improved by introducing heat into the system. However, this has been achieved by deteriorating the quality of the surroundings, for instance, by degrading a resource such as a fossil fuel that is burned to produce that heat. Having understood this, it becomes easier to link degradation to entropy. If an amount of heat, Q flows spontaneously from hot bodies (T) to cold ones (T0), with T > T0, then: _​ ​ ​  >  ​ _ ​(​ ​T​  ​) (​ T ​ )​ ​  Q

0

Q

f

i

(​ 31.3)​

The cold body has gained more entropy than the hot body has lost. The difference becomes greater as the temperature range between the two increases. This phenomenon is associated with the entropy generation as explained above (σ) and increases with the degradation of energy quality. Since all bodies tend to irreversibly dissipate their energy in the form 2 δQ of heat until reaching the temperature of their _ (​ 31.1)​ ​ ​∫ ​ ​  T ​ ) ​  ​  ​=  ( ​ ​  ​  ​∆ Srev surroundings, entropy will always be gener1 int rev ated. Only in the best possible case, known as where Q is the heat exchanged in the process the reversible process, is entropy generation and T is the temperature measured in Kelvin, zero, as expressed in Equation 31.1. at which the heat is transferred. This expresConsequently, in any process undertaken sion is derived from the well-known Clausius by a system, its entropy and that of the surinequality. Entropy has units of energy rounding environment will increase or, in the divided by temperature (i.e. Joules divided best case scenario, will remain constant. In by Kelvin in the International System of a mathematical way, the Second Law can be units (SI): J/K). Since the temperature is expressed as: always positive, but Q can be positive (heat transferred into the system) or negative (heat ​  >  0; with  ∆ ​Suniverse ​  ​ ∆ ​  ​  ​Suniverse dissipated by the system), the entropy of (​ 31.4)​ ​  ​ ​+  ∆ S​ surroundings ​  ​  ​​= ∆ ​Ssystem any internal reversible system undergoing a process can increase or decrease. The entropy change of non-ideal (irreversi- For an isolated system, meaning one in which neither matter nor energy can enter or exit the ble), or real systems, is expressed as: system, Equation 31.2 reduces to: ​​∆ Sirr  ​ ​ ​ ​+  σ ​ ​∫ ​ ​ _ ​  ​  ​=  ( T) 2 δQ

1

(​ 31.2)​​

where ​ σ​is called entropy generation and measures the irreversibilities associated with the process. The entropy generation is always positive and has the same units as property entropy S: J/K. In the same way as in a reversible process, the entropy change of a real system can be positive, negative, or zero, depending on the sign and magnitude of the δQ term ​∫1 2​ ​ _ T ​​ , which adds or counterbalances the

​  ​  ​=  ​+  σ ​∆ Siso

​(31.5)​

This means that, for an isolated system, the entropy can only increase with time. This fact is sometimes called “the arrow of time” (Lebowitz 1993). Hence, the Second Law of Thermodynamics can be formulated as follows: in all real energy transformation processes, the total entropy of all involved bodies can only be increased or unchanged in an ideal case. Beyond these conditions, the

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process is impossible, even if the First Law is fulfilled. The fundamental idea behind this is that everything degrades or becomes dispersed if there is no high-quality energy input into the system. So how does entropy relate to degradation? Unlike humans, nature is not purposeful in its actions. People design machines for domesticating nature’s forces and attempt to convert them into useful effects, as is the case of pumps used to revert the natural flow. Left alone, sooner or later, fluids will stop flowing, warm bodies will get colder, pure substances will become impure. Everything will degrade spontaneously. Degradation may thus be controlled and slowed to a certain extent, but it cannot be avoided in the long run. This is an experimental fact and constitutes the Second Law of Thermodynamics. So if everything degrades, yet energy is conserved, one is forced to admit that some forms of energy are more useful to society and less degraded than others. The quality of that energy with respect to a given reference environment is measured with another thermodynamic property called exergy,1 which unifies the First and Second Laws of Thermodynamics in a single numeraire. It is a fact that any form of noble energy (high-quality energy) will eventually degrade into heat at the lowest possible temperature. All other types of energy will arrive at this point sooner or later. This is why the Second Law is so devastating—it truly announces dissipation. Thermodynamics does not state when this will happen but does guarantee that it will happen. Recycling processes are also all subject to the Second Law. So, even if the aim of recycling is the recovery of materials from waste and, in doing so, apparently decreasing entropy, in each cycle, some quantity and quality of materials is unavoidably lost (strictly speaking, there are no circles, but spirals; Valero et al. 2021). This is something that Georgescu-Roegen recognised and even formulated in his “Fourth Law” (Georgescu-Roegen 1971), claiming that matter behaves in a way parallel to the fact that available energy is irrevocably converted into an unavailable one. Thus, in his opinion, available matter irrevocably becomes unavailable. The above explanations make it easier to understand what is referred to in ecological

economics by the terms “high entropy” or “low entropy” when describing the world. Low-entropy examples include natural feedstock, fossil fuels, food, wood, a lake, or even the sun. Helped with a spark, a fossil fuel reacts with oxygen in the air, transforming its energy into heat. Food is metabolised in a controlled combustion reaction, maintaining body temperature constant and providing energy to meet its needs. Later, that same energy dissipates and heats the environment. Thermal energy is an unimaginably huge energy drain. This simple fact allows for a better understanding of nature. Therefore, all physical activities are linked to irreversibility, and the greater the irreversibility, the faster the arrow of time shortens. On a finite planet, any activity, however minor, is connected to entropic degradation through climate change, depletion of non-renewable resources, or the alteration and displacement of the natural environment. All are subject to the Second Law. These ideas reflect another: entropy production or destruction does not always have negative consequences, as seen in the beauty of a river’s waterfall. Nature does not attempt to restrict the Second Law machine. On the contrary, humans installing a hydraulic turbine avoid entropy generation at the cost of the river’s aesthetics. It is worth noting that the greatest entropy production taking place on Earth is not due to anthropogenic activity but is solar in origin. The sun radiates heat towards the Earth at an average temperature of around 15°C, at approximately 5500°C (i.e. the sun’s energy is 95 per cent pure exergy). Hence, the Earth is not an isolated system. Waves, winds, the formation of clouds, and the heating of the Earth are only intermediate manifestations of that dissipation. Perhaps surprisingly, only 0.023 per cent of the entire solar energy is stored in the form of biomass. Yet, with this tiniest amount, the Earth has been able throughout time to steadily store all the fossil fuels that people have burned in the last few generations. This gives thermodynamic hopes to the future since, by avoiding a small part of that degradation (i.e. using technology that could take advantage of solar energy in all its forms), people could stop or at least slow down the depletion of the planet. Speaking about conserving nature often infers conservation of all natural energy dissipation processes, including rivers full of

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freshwater irreversibly mixing with seawater in estuaries, air flows cooling warm areas, and even tectonic activity occurring with some frequency. Maintaining the delicate equilibrium of life cycles and planet diversity is subtler than performing mere entropic balances. However, the Second Law is present in all natural processes and, when analysing why they occur, Second Law-thinking helps. Ironically, it is life while appearing completely anti-entropic, not death, that dissipates energy. Societies, as living entities, behave similarly. It is important to clearly understand this phenomenon if humankind as a mere pattern of the Second Law is ever going to decelerate the passing of time. Only intelligence through good housekeeping will prolong life. Alicia Valero, Antonio Valero, and Guiomar Calvo

Note

1. Exergy is capable of simultaneously expressing both the quantity and quality of an energy flow. While energy is conserved and transferred, exergy always degrades. With exergy, not only energy sources can be assessed, but also any natural resource, once the physical properties of the system and the conditions of the surroundings (or reference environment) are set (Szargut 2005). Accordingly, the quality and quantity of resources such as minerals, water or even soils and their degradation velocity can be evaluated with a single nummeraire measured in energy units (Lebowitz 1993).

References

Georgescu-Roegen, N. (1971). The Entropy Law and the Economic process. Harvard University Press, Cambridge, MA. Lebowitz, J.L. (1993). Boltzmann’s Entropy and Time’s Arrow, Physics Today, 46, 32–8. Szargut, J. (2005). Exergy Method: Technical and Ecological Applications. WIT Press, Southhampton, Boston. Valero, A., Valero, A. and Calvo, G. (2021). The Material Limits of Energy Transition: Thanatia. Springer, Cham.

Alicia Valero, Antonio Valero, and Guiomar Calvo

32. Environmental accounting 32.1 Introduction

Concern about the impact of human activities on the environment has been raised most often, for example, in issues regarding public health, air, and water pollution (to name a few) and, more recently, biodiversity. Historically, the birth of environmental accounting cannot be traced back to a single researcher or date as it is inspired by several schools of thought and the long-held debates on sustainability and the (in)compatibility between unlimited growth and the environment. In the 1960s, a few heterodox economists and ecologists (including Nicholas Georgescu-Roegen, Kenneth Boulding, K.W. Kapp, Herman Daly, and H.T. Odum) began to look at the economy as a subsystem embedded in a physical system of materials and energy, and started to conceptualise and quantify the effects of human activity on the environment to track and measure progresses towards sustainability outcomes and goals. Fifty years after The Limits to Growth book (Meadows et al., 1972), which raised awareness of our planet’s limited resources, scientists and policy makers are still working to assess both direct and indirect impacts of human activity (Patterson et al., 2017). This entry showcases the broad spectrum of diverse approaches to environmental accounting that have developed during the last 50 years. Although the expression ‘environmental accounting’ refers to a variety of metrics and methods, this entry restricts discussion to ‘biophysical’ methods, which primarily use biophysical metrics. Therefore, this entry does not focus on integrated national accounts on macro-economic activity and the environment, nor does it cover the integration of environmental data into business and firm-level financial accounting and auditing, most often also defined as environmental accounting methods. This entry is not intended to present a comprehensive review of all the environmental accounting methods (which are deeply discussed in dedicated entries of this encyclopedia), and it does not provide a systematic comparison of their features, which can be

found elsewhere (Bakshi, 2019; Nicolucci et al., 2001; Kharrazi et al., 2014; Patterson et al., 2017; Rodríguez et al., 2019). Rather, it aims to help readers navigate the complexity and interconnectedness of environmental accounting methods for deeper understanding and synergic use. Environmental accounting methods can be classified in several ways, for example, depending on if they primarily look at upstream or downstream impacts, which has limited validity when accounting methods encompass several indicators, and/or depending on the research aim. For this reason, this entry proposes to group the different methods in three categories (thermodynamic-oriented, socio-metabolic oriented, and impact-oriented) depending on the guiding principles and scoping of the different approaches.

32.2 Thermodynamic-oriented methods

The idea that the control of energy matters for society and even determines the advancement of civilisation moved researchers to focus on the energy analysis of systems (Smil, 2010). Basically, the Laws of Thermodynamics are relevant to the economy because economic activity is entropic: the economic system is a dissipative structure that does not produce but rather converts available resources into degraded energy and matter in support of its functioning. Thermodynamic concepts have been utilised by practitioners in a variety of disciplines with interests in environmental sustainability, including ecology, economics, and engineering. It has been argued that these consequences of human development are reflected in thermodynamic ideas and methods of analysis; they are said to mirror energy transformations within society. 32.2.1 Energy and exergy analysis After the energy crisis of the ’70s, Net Energy Analysis (NEA) has been employed as a structured, comprehensive method of quantifying the extent to which a given energy source provides an energy surplus to end users, after accounting for energy losses occurring along the chain of processes to exploit it (i.e. extraction, refining, transformation into a usable energy carrier, and delivery to end users), thanks to the efforts of several crea-

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Note: Environmental accounting methods categorisation according to their guiding principles and scoping. Examples of integration among methods are displayed by grayscale lines depending on the level at which the integration occurred: pre-analytical stage (light gray line), analytical stage (black line), and post-analytical stage (dark gray line). While some methods have been integrated more often in the literature (e.g. Life Cycle Assessment and Emergy Accounting [LCA-EMA]), others have remained more isolated, although they cannot be excluded, as they used data inventories from Material and Energy Flow Analyses (MEFA) and LCA, for example.

Figure 32.1 Graphical abstract

tive and broad-minded scientists (Malcom Slesser, Howard T. Odum, Charlie Hall, Robert Herendeen, among others). Energy Return on (energy) Investment (EROI; see also a more detailed description in this book, at Chapter 29 by Charlie Hall and Rigo Melgar), the ratio of the energy delivered by a process to the energy used directly and indirectly in that process, is the most common metric of NEA and it has been used to evaluate and quantify the net or surplus energy eventually used or gained for a given project, industry, nation, fuel, or resource (Hall et al., 1986). Therefore, the focus is not on the total amount of primary energy used (as in the impact-oriented methods), but on the energy gained by society. However, the energy analysis does not always provide sufficient insight in that it focuses on the amount of heat potentially available from a source. More advanced analysis methods, such as exergy analysis, can be applied to provide a deeper understanding of the usefulness of an energy source in a specific process Maddalena Ripa and Sergio Ulgiati

(a measure of user-side energy quality or work potential, namely, ability to do work). Exergy is a concept derived from the First and Second Laws of Thermodynamics (Szargut et al., 1988; Szargut, 1989). The main important difference between energy and exergy is that the former is conserved, while the latter can be consumed. Therefore, exergy provides a measure of how nearly the actual performance of a system approaches the ideal, and it identifies the causes and locations of thermodynamic losses much more clearly than energy analysis. Jørgensen et al. (2007) has proposed a modified approach to exergy for specific usage in an ecological context: eco-exergy. This measures the informational complexity of systems through the content of DNA. Energy analysis approaches have evolved from the previous studies in the 1970s and 1980s and, although much work has been carried out to estimate the EROI of individual renewable energy source technologies, important methodological discrepancies still

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exist depending on the applied definition, the system design and location, the functional units, or the boundaries of the analysis. While there have been some attempts to standardise these approaches, the lack of methodological consistency has led to a situation where inappropriate comparisons are being made across technologies and among (energy) carriers that cannot be put to similar end-use, thus hindering the potential to use the EROI concept to adequately inform policy makers and society as a whole (Raugei, 2019; Murphy et al., 2022). 32.2.2 Emergy accounting Emergy accounting (EMA; Odum, 1988, 1996) is an accounting method aimed at assessing the environmental performance and sustainability of processes and systems on the global scale of the biosphere (see also a more detailed description in this book, at Chapter 28 by Silvio Viglia and Sergio Ulgiati). It takes into account free environmental inputs (e.g. solar radiation, wind, rain, and geothermal flows), human-driven flows, and the indirect environmental support embodied in human labour and services. In EMA, all inputs supporting a system are accounted for in terms of their ‘solar emergy’, defined as the total amount of solar available energy directly or indirectly required to make a given product or to support a given flow or service (i.e. the work performed by nature over the entire supply chain), and expressed by means of one energy form only (generally the solar form, measured as solar emergy joules, sej). Compared to other environmental analysis methods, EMA carries the added value of unifying flows of different natures by means of its unite emergy value (UEV) quality factors, thus capturing a comprehensive eco-centric value of products and processes. Nevertheless, like many ground-breaking ideas, EMA has encountered resistance and criticism, particularly from economists, physicists, and engineers, since its early years: the initial lack of standardisation and the apparent disregarding of the fundamental principles of market demand and mainstream economics (limited database, conflicting relation to market dynamic, different concept of value) were the major critiques to the method (Månsson and McGlade, 1993).

32.3 Socio-metabolic-oriented methods

Metabolism is a biological concept that describes the chemical conversion of materials and energy by organisms to sustain growth and reproduction. Social metabolism by analogy implies that society—similar to an organism—supports its structures (in-use stocks and funds) through exchanging energy and materials with the surrounding environment (Yan et al., 2020; i.e. all the energy and material transformations that occur within an open social system, such as an economy, and between this system and its environment; Gerber and Scheidel, 2018). These complex processes determine the functional structure of the system, ensure its reproduction, maintain and repair its parts, and exhibit specific dynamics according to different contexts, as described in the following section. In other words, social metabolism can be used to measure the process by which a society transforms energy and matter to ensure its continued existence. Next, two methodological approaches are briefly presented; for more information regarding a detailed comparison between the two, see Gerber and Scheidel (2018). 32.3.1 Material and energy flows analyses MEFA is based on the notion of socioeconomic metabolism (Fischer-Kowalski and Hüttler, 1998; Haberl et al., 2019; Schmidt-Bleek, 1992; see a more detailed description in this book in Chapter 61 by Fridolin Krausmann). MEFA takes into account all the materials and energy used by national economies, including water, air, and fossil fuels. The ultimate goal of the analysis is drawing a picture of the biophysical patterns and dynamics of socioeconomic material and energy flows as well as their underlying drivers. MEFA has been used to analyse the metabolisms of large-scale systems, such as countries or regions, and to link them to broad historical analyses, such as long-term socioecological transitions as well as in the context of prospective models and scenarios of future resource use (Krausmann et al., 2017). Some of the future research strands in MEFA involve advancement in: (i) the disaggregation of different qualities of material flows to derive (more robust) aggregated Maddalena Ripa and Sergio Ulgiati

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indicators; (ii) the links between MEFA indicators and environmental impacts, assessing indirect material flows of traded products (implying possible shifts of environmental burden); and (iii) the comprehensive output accounts (historically less advanced and not yet standardised; Krausmann et al., 2017). 32.3.2 Multi-scale integrated analyses of societal and ecosystem metabolism The multi-scale integrated analysis of societal and ecosystem metabolism (MuSIASEM method; Giampietro et al., 2009) has been in development since the 1990s (see a more detailed description in this book, at Chapter 66 by Mario Giampietro). It builds on Georgescu-Roegen’s (1971) bioeconomics concepts and complex systems theory (Allen and Starr, 1982; Rosen, 1991). The proponents of MuSIASEM argue that since socioecological systems are self-organised, the study of metabolism requires considering their hierarchically organised structural and functional compartments operating at different space dimensions and multiple scales. Specifically, following the hierarchy theory, MuSIASEM connects the behaviour of a system as a whole (e.g. a region) to the behaviour of its functional parts (e.g. its economic sectors) and makes the distinction between funds—system elements that process, metabolise, or transform material flows (e.g. capital, people, or land)—and flows—system elements that are produced, consumed, and transformed (e.g. energy, food, or water). MuSIASEM is commonly used to analyse the societal metabolism of countries, sectors, and energy systems, among others. By combining extensive and intensive variables derived from socioeconomic, demographic, and energy data, MuSIASEM allows us to investigate how changes taking place in a specific dimension and scale can transform/ constrain the entire system together with the composing parts. In MuSIASEM, a system is assessed in terms of feasibility (compatibility with processes outside human control), viability (compatibility with processes under human control), and desirability (compatibility with normative values and institutions). Although most of the papers have been historically focused on the feasibility and viability dimensions with desirability set aside, the recently updated version of the MuSIASEM, supported by quantitative Maddalena Ripa and Sergio Ulgiati

storytelling (Matthews et al., 2017), has made a remarkable effort to represent alternative development perspectives while still leaving a marginal place for dialectical reasoning (Beltramello and Bootz, 2022).

32.4

Impact-oriented methods

Loosely, the following methods try to answer to the question: ‘What is the actual or potential impact of a given product/process/service?’ To describe and measure the load of human society on ecological systems, scientists have long strived to develop comprehensive indicators, of which environmental footprints are probably the most popularly recognised and employed. As a matter of fact, some of the footprint indicators (such as land use, water use, greenhouse gas [GHG] emissions) are being less used as they represent a sub-set of the data covered by a more complete LCA). Nevertheless, although it is commonly assumed that with footprint analyses LCA is ‘streamlined’ to cover solely one indicator, we discuss the main discrepancies in their conceptual and practical applications (Fang et al. 2013; Fang and Heijungs, 2014). 32.4.1 Footprint ‘family’ Unlike the above energy, exergy, and emergy approaches, which are rooted in complex thermodynamics and ecology concepts, the footprint family is based on much simpler, although very telling, definitions that may explain their popularity and frequent use. Although originally ‘footprint’ meant, quite literally, a land area appropriated by some entity—its imprint on the land (Wackernagel and Rees, 1996)—later the general definition was extended to ‘indicators of human pressure on the environment’ (Hoekstra and Wiedmann, 2014). In this entry, the most common environmental footprints (i.e. the so-called ‘footprint family’) will be reviewed, namely, the ecological, carbon, and water footprints. The conceptual basis of a footprint is that everything produced, consumed, or utilised by humans requires inputs from the natural world (usually as resources) or generates outputs (carbon emissions) that can be quantified. As such, footprints reflect the embedded resources in products that, when traded or sold, constitute a virtual exchange of them. Virtual or embedded resources in

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products/services can be imported, exported, or utilised locally. 32.4.1.1 Ecological footprint An ‘ecological footprint’ (EF; see a more detailed description in this book, at Chapter 51, The human ecological footprint by William E. Rees) is a cumulative surface-based measure of the pressure that humans exert on the planet. The EF approach compares human demands on nature with the biosphere’s ability to regenerate resources (Wackernagel and Rees, 1996). EF can be calculated for an individual, a village, a locality, a nation, or the whole world itself, along with organisations, services, products, and processes, expressing impacts in terms of global hectares (gha) or ‘number of planets’ directly and indirectly needed to support a process, an individual, or a system. Despite its rapid ascent and widespread use, the EF has faced a wide range of criticism, among which are its provision of a single aggregate indicator of impacts and its failure to capture several important human pressures on the Earth systems, for example, the decline of biodiversity, eutrophication, consumption of mineral resources, or emission of toxic compounds (Galli et al., 2016; Giampietro and Saltelli, 2014). 32.4.1.2 Carbon footprint The term ‘carbon footprint’ (CF) describes a large variety of methodological approaches that assess the impact of an activity on the Earth’s climate (Wiedmann and Minx, 2008) by quantifying carbon emissions (CO2, CH4, and other GHGs) in terms of CO2 equivalents (CO2-eq). The concept is ‘catchy’ and has been promoted and diffused outside the research community. That being said, relying entirely on one indicator can sometimes be misleading (that is, problem-shifting, solving one environmental problem but likely disregarding others in the process). 32.4.1.3 Water footprint The water footprint (WF) quantifies the amount of freshwater used for a product over the entire supply chain (see also a more detailed description in this book, at Chapter 90 by Cristina Madrid-López). WF accounts for both direct (domestic water use) and indirect (water required to produce industrial

and agricultural products) water uses and has become an important indicator to track human pressure on freshwater sources. The strength of the WF concept is that it allows a broad perspective on the water management of the system and a deeper understanding of water usage. Its main weaknesses are that it represents just the quantity of water used (without an estimation of the associated environmental impacts), the lack of required data, and the subjectivity of grey WF estimates (Jeswani and Azapagic, 2011). Two main approaches to WF exist in the literature: one developed by the Water Footprint Network (WFN; Mekonnen and Hoekstra, 2011) and the LCA approach developed by the LCA community (which includes a weighted WF approach). The primary difference between WF and LCA is that the former focuses on the volume of water and the latter on the impacts associated to water use. Indicators of water usage and impacts derived from the two methods are potentially compatible, but are not directly comparable (Fresán et al., 2019; Hoekstra, 2016; Pfister et al., 2017) 32.4.2 Life cycle assessment LCA is a methodological framework for assessing potential environmental loads and uses of resources in human-dominated processes, depicted in details by International Organization for Standardization (ISO, 2006a, 2006b) standards and by the International Reference Life Cycle Data System (ILCD) Handbook from Joint Research Centre (JRC, 2010), providing a ‘cradle to grave’ approach, from resources extraction to disposal of waste. LCA is standardised into four main phases: i) goal and scope definition; ii) inventory analysis; iii) impact assessment; and iv) interpretation. ● Goal and scope definition: stating the reason behind the study, the intended application, the audience and whether the results are intended to be disclosed to the public or not, the choice of a reference item for calculations (functional unit), and system boundaries. ● Life cycle inventory analysis: input and output flows are listed and collected for each step (unit process) and properly referred to time, geography, and source. Background data for the investigated supply chain inventory can be retrieved Maddalena Ripa and Sergio Ulgiati

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from specialised commercial (e.g. https://​ecoinvent​.org/​) and open access databases. ● Life cycle impact assessment (LCIA): represents the translation, by means of several impact assessment methods, of the built inventory of input and output flows into environmental impacts, expressed as a set of indicators (e.g. global warming, eutrophication, acidification, etc.) representing the potential burdens happening at the technosphere–ecosphere interaction. ● Interpretation: final results are interpreted according to the initially defined goal and scope, inventory, and boundary. There are two types of LCAs discussed widely in the literature, namely, attributional LCA (ALCA), attributing a defined allocation of environmental impacts to a product or process unit, and consequential LCA (CLCA), which considers wider system effects of change. The main applications of LCA focused on the analysis of the contribution of the life cycle stages to the overall environmental load, usually with the aim to prioritise improvements on products or processes and to make comparisons between products/ services. Despite the numerous studies and publications, LCA also presents some inherent limitations related to complex modelling, blurred or inconsistent boundary setting, short time horizons during which the outputs are considered to be valid, budget, data availability and quality, and the lack of consideration of geographical and local sensibilities to environmental impacts (Guinée et al., 2011; Guinée and Heijungs, 2011). A final topic of concern is the translation from functional-unit-based to real-world improvements, including the fact that LCA does not sufficiently represent the social and economic interactions between key actors in the analysis (Gutowski, 2018). Many efforts are being carried out, both at conceptual and practical levels, including the development of regionalised databases, new impact assessment methods and uncertainty analyses, as well as the tentative broadening of the scope of LCA by including life cycle costing (LCC) and social LCA (SLCA), thus covering all three dimensions of sustainability (i.e. people, planet, and prosperity).

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32.5

The final aim of environmental accounting: delivering more than the sum of the parts

To overcome the shortcomings of each methodology and develop more holistic assessments, the integration of these methodologies is essential (Ulgiati et al., 2006; Oliveira et al., 2021). In recent years there has been an increasing trend to ‘integrate’ or ‘use alongside’ each other the different environmental accounting methods to gain a more complete picture of the environmental impact, rather than relying on one method that usually has a single criterion perspective. Several studies have attempted to integrate these methodologies either conceptually or through applied case studies. The integration can occur at different levels in an effort to yield more comprehensive inventories and assessments and increasing consistency: in the pre-analytical step, two or more methods use (partially or fully) the same data inventory or a combination of them, but analyses are conducted independently from each other; in the analytical step, two or more methods are used in a sequential analytical procedure; in the post-analytical step, methods are conducted independently from each other but their results are combined/aggregated by introducing a qualitative interpretation of the result. A telling example of the first category is Brunner and Rechberger (2004): an LCA inventory is integrated with economy-wide MFA to broaden the scope of the analysis and for application to different systems (e.g. economic sectors or cities) and different research questions (e.g. for assessing environmental impacts related to material flows or to analyse flows within the economy and recycling). In this same category, Lyu et al. (2021) provide an interesting case dealing with LCA and emergy assessment of agricultural chemicals production, based on the same inventory from the Ecoinvent database. The allocation default of the database for LCA use is reversed to the original data without allocation to meet the emergy algebra requirements and to calculate the conversion factors (UEV) needed for the emergy evaluation of intermediate and final chemicals. A very clear example of the second category is Santagata et al. (2020), dealing with alternative patterns for paper production.

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An ex ante LCA of the business-as-usual process identifies its main hotspots, followed by the selection of potential alternatives and a sequential application of the EMA to each option to assess their large-scale environmental sustainability (and potentially discarding some of them). Finally, an ex post LCA for each alternative allows us to ascertain if the hot spots are likely to be removed by the selected alternative process. The third category is clearly exemplified by Gasparatos et al. (2009a, 2009b) where MuSIASEM and EMA results are (dialectically) combined: MuSIASEM is used to elucidate the shift of the UK economy from an industrial to a service economy in the period between 1981 and 2004, while EMA is employed to understand the production and consumption patterns and the environmental support required to sustain human activity within the UK for the year 2004. While some methods have more often been integrated in the literature (e.g. LCA-EMA), others have remained more isolated, although they cannot be excluded, as they used data inventories from MEFA and LCA, for example. To conclude, the notion that environmental problems can be dealt with in individual silos is long gone and so should be for environmental accounting. Maddalena Ripa and Sergio Ulgiati

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Environmental accounting  199 Emergy Analysis, Life Cycle Assessment and Other Methods of Environmental Accounting. Ecological Modelling 362, 19–36. https://​ doi​ .org/​10​.1016/​J​.ECOLMODEL​.2017​.07​.022 Pfister, S., Boulay, A.M., Berger, M., Hadjikakou, M., Motoshita, M., Hess, T., Ridoutt, B., et al., 2017. Understanding the LCA and ISO Water Footprint: A Response to Hoekstra (2016) ‘A Critique on the Water-Scarcity Weighted Water Footprint in LCA’. Ecological Indicators 72, 352–9. https://​doi​.org/​10​.1016/​J​.ECOLIND​ .2016​.07​.051 Raugei, M., 2019. Net Energy Analysis Must Not Compare Apples and Oranges. Nature Energy 4, 86–8. https://​doi​.org/​10​.1038/​S41560​-019​ -0327​-0 Rodríguez, C.M., Rengifo Rodas, C.F., Corrales Muñoz, J.C., Casas, A.F., 2019. A Multi-Criteria Approach for Comparison of Environmental Assessment Methods in the Analysis of the Energy Efficiency in Agricultural Production Systems. Journal of Cleaner Production 228, 1464–71. https://​doi​.org/​10​.1016/​J​.JCLEPRO​ .2019​.04​.388 Rosen, R., 1991. Life Itself: A Comprehensive Inquiry into the Nature, Origin, and Fabrication of Life. Columbia University Press. Santagata, R., Zucaro, A., Fiorentino, G., Lucagnano, E., Ulgiati, S., 2020. Developing a Procedure for the Integration of Life Cycle Assessment and Emergy Accounting Approaches. The Amalfi Paper Case Study.

Ecological Indicators 117, 106676. https://​doi​ .org/​10​.1016/​j​.ecolind​.2020​.106676 Schmidt-Bleek, F., 1992. MIPS – A Universal Ecological Measure. Fresenius Environmental Bulletin 2, 407–12. Smil, V., 2010. Energy Transitions: History, Requirements, Prospects. Praeger. Szargut, J., 1989. Chemical Exergies of the Elements. Applied Energy 32, 269–86. https://​ doi​.org/​10​.1016/​0306​-2619(89)90016​-0 Szargut, J., Morris, D., Steward, R., 1988. Exergy Analysis of Thermal, Chemical, and Metallurgical Processes. Hemisphere Publishing Corporation. Ulgiati, S., Raugei, M., Bargigli, S., 2006. Overcoming the Inadequacy of Single-Criterion Approaches to Life Cycle Assessment. Ecological Modelling 190, 432–42. https://​doi​ .org/​10​.1016/​J​.ECOLMODEL​.2005​.03​.022 Wackernagel, M., Rees, W.E., 1996. Our Ecological Footprint. Reducing Human Impacts on Earth. New Society Publications. Wiedmann, T., Minx, J., 2008. A Definition of ‘Carbon Footprint’. In: C. C. Pertsova, ed., Ecological Economics Research Trends, 1–11. Nova Science Publishers. Yan, N., Liu, G., Ripa, M., Wang, N., Zheng, H., Gonella, F., 2020. From Local to National Metabolism: A Review and a Scale-Up Framework. Ecosystem Health and Sustainability 6, 1839358. https://​doi​.org/​10​ .1080/​20964129​.2020​.1839358

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33. The environmental consequences of inequality Inequalities of wealth and power are implicated in both the distribution of environmental costs and their total magnitude. Activities that release pollutants and deplete natural resources result in winners as well as losers. Some actors benefit (or at least think they do) from these activities; otherwise, the activities would not occur. Others bear net costs, suffering environmental harms that outweigh whatever benefits, if any, they obtain. The ability of winners to reap benefits for themselves while imposing costs on others is a function of their relative economic and political power. Purchasing power underpins effective demand in markets: willingness to pay exists if, and only if, preferences are backed by the ability to pay. Every dollar counts equally in the market, so those with more dollars wield more “votes.” Willingness and ability to pay likewise underpin the techniques used in cost–benefit analysis to assign monetary values to non-market goods, such as clean air, clean water, and a stable climate. Insofar as environmental policy decisions are dictated by the efficiency criterion of neoclassical economics – weighing benefits against costs to determine the “optimal” levels of pollution and natural resource depletion – those with more purchasing power again wield more votes. Political power reinforces and magnifies the effects of purchasing power. When those who are harmed by environmental degradation lack political influence – for example, when they are unable to vote or to lobby elected officials – the costs and benefits that accrue to them carry less weight in social decisions. In the extreme case where those who are harmed have no political power whatsoever, decision-makers can simply ignore the costs imposed upon them. An example is the US Environmental Protection Agency’s 2017 decision, in analyzing the costs and benefits of curbing emissions from electric power plants, to assign zero value to climate damages incurred outside the United States, a stance that put an official imprimatur on the maxim “out of sight, out of mind.”

Purchasing power and political power are mutually reinforcing: those with more wealth typically wield more political influence, and vice versa. The joint effect of these two dimensions of power can be characterized by a power-weighted social decision rule (PWSDR), in which environmental outcomes maximize net benefits weighted by the political influence of those to whom they accrue (Boyce 1994). Two predictions follow. The first is that the distribution of environmental costs will not be random. Instead, risks and harm are likely to be inflicted disproportionately on those with less wealth and power. The second is that wider inequalities of wealth and power will result in higher levels of environmental degradation overall. Both propositions have been supported in recent years by a growing body of empirical research.

Inequality and the distribution of environmental costs

The logic behind the first prediction is straightforward. Environmental quality generally is not a pure public good that when available to one is equally available to all. Rather it is an impure public good; in George Orwell’s memorable phrase, “Some are more equal than others.” Likewise, environmental degradation is an impure “public bad.” Once we recognize that environmental costs are not impersonal misfortunes that fall indiscriminately across the population, we can expect to see them imposed disproportionately on individuals, communities, and nations that lack the economic and political power to fend them off. Pollution has particularly adverse effects on children, leading to higher infant mortality, lower birthweights, a higher incidence of neurodevelopmental disabilities, more frequent and severe asthma, and lower school test scores (e.g., Currie 2011). Among adults, pollution exposure leads to higher morbidity and mortality and to more lost work days due to illness and caring for sick children (Boyce et al. 2016). These impacts exacerbate the vulnerabilities that make some communities more susceptible to environmental harm in the first place. In the United States, environmental justice researchers have documented systematic disparities in exposure to pollution and other hazards along the social fault lines

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of race, ethnicity, and income. A pioneering study by sociologist Robert Bullard (1983) showed that hazardous waste disposal sites in Houston, Texas, were sited primarily in African-American neighborhoods. Subsequent research identified similar patterns across the country. Race and ethnicity often are stronger than income as predictors of proximity to hazards and pollution exposure, testifying to their salience in the distribution of political power (Zwickl et al. 2014). Investigations of the causal linkages that underlie these spatial correlations have found clear evidence of disparities in the initial siting decisions, as well as some evidence of post-siting demographic shifts (Mohai and Saha 2015). In Delhi, India, where residents breathe some the world’s dirtiest air, researchers similarly have found that the poor generally live in more polluted neighborhoods and spend more time working outdoors where pollution exposures are most intense (Garg 2011; Foster and Kumar 2011). Inequalities based on gender often translate into disparate environmental harms inflicted on women. A prime example is the disproportionate exposure of women to indoor air pollution in south Asia and sub-Saharan Africa, where solid fuels such as wood, crop residues, and dung are widely used for cooking (Agarwal 2010; Okello et al. 2018). The World Health Organization (WHO 2021) estimates that this pollution is responsible for more than 3.8 million deaths per year, mostly of women and children. Environmental inequalities also extend across national borders. In an infamous 1992 memorandum, the World Bank’s chief economist argued that “the economic logic of dumping a load of toxic waste in the lowest-wage country is impeccable” (Summers 1992, 66). The statement was said to have been meant provocatively, but real-world practice often follows this script. Each year millions of tons of toxic waste are shipped from the industrialized nations of the Global North to lower-income countries in Africa, Asia, and Latin America. The Basel Convention on the Control of Transboundary Movement of Hazardous Wastes and their Disposal, an international environmental accord that went into effect in the same year as the World Bank memo, has proven inadequate to prevent the large-scale shifting of

environmental costs onto some of the world’s poorest people.

Inequality and the magnitude of environmental degradation

The second prediction following from the PWSDR is that wider inequalities in the distribution of wealth and power will tend to result in more environmental degradation overall. The concentration of environmental harms at the lower end of the wealth-and-power spectrum implies that with wider inequality, such costs carry less weight on both the economic scales of markets and the political scales of policy makers. Moreover, the benefits from environmentally degrading activities tend to be concentrated at the upper end of the wealth-and-power spectrum. Externalization of environmental costs yields higher profits for shareholders, higher compensation for the firm’s executives, and sometimes lower prices for its consumers. Shareholders and executives typically occupy high rungs on the spectrum, and benefits passed through to consumers accrue disproportionately to those with the most purchasing power. With wider inequality, all these benefits carry greater weight both in markets and in the eyes of policy makers. In other words, both the cost side and the benefit side of the scales are tipped in favor of more pollution and resource depletion. From the perspective of methodological individualism – the foundational perspective of neoclassical economics – in which social outcomes are reduced to the sum of individual preferences weighted by ability to pay, wider income inequality might be expected to lead to less environmental degradation rather than more, insofar as the share of household expenditure devoted to energy and other resource-intensive goods tends to decline as incomes rise (Boyce 2007). From the perspective of political economy, however, interrelationships among people matter as well as individual preferences. Economies, like ecosystems, are not merely the sum of their individual parts: they are societal webs of interactions and interdependence. Many affluent people prefer to live in a clean and safe environment. But because environmental quality is an impure public good – neither entirely private nor equally available or unavailable to all – they can James K. Boyce

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reside in less contaminated locations. They also can afford to shield themselves from the impacts of pollution by buying air purifiers and drinking bottled water. In the event of pollution-related illness they can obtain better medical care. And they can deploy their political leverage to oppose the siting of environmental hazards in their neighborhoods, and to demand more stringent regulation of hazards they cannot avoid. The wealthy and powerful cannot escape the consequences of environmental degradation altogether, but in their private calculations they weigh a relatively small share of the costs against a relatively large share of the benefits. A number of empirical investigations have supported the prediction that wider inequality will result in more environmental degradation. When cross-national data on environmental variables first became available in the 1990s, early studies on how pollution varies with per capita income found that in many cases there was evidence of an inverted U-shaped relationship that came to be known as the environmental Kuznets curve (Grossman and Krueger 1995). When measures of economic and political inequality were added to the analysis, it was found that higher levels of inequality are associated with more pollution, and that the apparent relationship between pollution and per capita income often weakens or disappears once the impact of inequality is taken into account (Torras and Boyce 1998; Farzin and Bond 2006). Similarly, researchers have found that biodiversity losses are more severe in countries where income is more unequally distributed (Mikkelson et al. 2007; Holland et al. 2010). In general, evidence for the adverse environmental effects of inequality is strongest for harms with immediate and visible impacts on human health, as one might expect (Cushing et al. 2015). For impacts such as climate destabilization that are more widely dispersed across time and space, the evidence is less conclusive. Nevertheless, several recent studies have found a correlation between inequality and carbon dioxide emissions (Knight et al. 2017; McGee and Greiner 2018). One explanation may be that fossil fuel combustion releases hazardous “co-pollutants” with immediate and localized effects, alongside carbon dioxide, James K. Boyce

and that this helps to spur localized public demand for emissions reductions. Within the United States, researchers have found evidence that states with higher levels of inequality tend to have more severe environmental degradation. More unequal distribution of power at the state level is associated with weaker environmental policies, more environmental stress, and worse public health outcomes (Boyce et al. 1999). Inter-state differences in inequality have also been found to be correlated with carbon dioxide emissions (Jorgenson et al. 2017). Within China, cross-sectional analysis similarly has found that regions with higher income inequality tend to suffer worse air pollution (Wang et al. 2021). Inequality can also foster environmental degradation by eroding concern for the well-being of future generations. For the rulers in highly unequal societies, the risk that their political power may one day come to an end encourages a cut-and-run approach to natural resource extraction, as exemplified in the rapacious deforestation experienced in Southeast Asia under the dictatorships of Ferdinand Marcos in the Philippines and Suharto in Indonesia. At the same time, for the poorest people the demands of day-to-day survival may override worries about tomorrow. The latter effect can also be seen in settings with less extreme inequality. After the French government’s 2018 announcement of a fuel tax increase to combat climate change, Yellow Vest protestors took to the streets, explaining that President Macron “talks about the end of the world while we are talking about the end of the month” (Rubin 2018; see also Mehleb et al. 2021). Combining these two environmental predictions, the first on the distribution of environmental costs and the second on their magnitude, we can expect that locations with wider environmental disparities will tend to have more environmental degradation overall. In line with this, researchers in the United States have found that all population groups suffer more severe air pollution and higher cancer risks in metropolitan areas that have a higher degree of residential segregation and wider racial and ethnic disparities in pollution exposure (Morello-Frosch and Jesdale 2006; Ash et al. 2013). In sum, then, our relationships with the environment are closely intertwined with our

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relationships with each other. Environmental degradation is not simply a matter of humans harming other species and ecosystems; it is also a matter of some people harming other people. To rebalance our relationships with nature, it will be necessary to rebalance our relationships among ourselves. James K. Boyce

References

Agarwal, B. 2010. Gender and Green Governance. Oxford: Oxford University Press. Ash, M., Boyce, J.K., Chang, G., and Scharber, H. 2013. Is environmental justice good for white folks? Social Science Quarterly 94, 616–36. Boyce, J.K. 1994. Inequality as a cause of environmental degradation. Ecological Economics 11, 169–78. Boyce, J.K. 2007. Inequality and environmental protection. In Jean-Marie Baland, Pranab Bardhan, and Samuel Bowles, eds., Inequality, Collective Action, and Environmental Sustainability. Princeton, NJ: Princeton University Press, 314–48. Boyce, J.K., Klemer, A.R., Templet, P.H., and Willis, C.E. 1999. Power distribution, the environment, and public health: A state-level analysis. Ecological Economics 29, 127–40. Boyce, J.K., Zwickl, K., and Ash, M. 2016. Measuring environmental inequality. Ecological Economics 124, 114–23. Bullard, R.D. 1983. Solid waste sites and the Black Houston community. Sociological Inquiry 53, 273–88. Currie, J. 2011. Inequality at birth: Some causes and consequences. American Economic Review Papers & Proceedings 101, 1–22. Cushing, L., Morello-Frosch, R., Wander, M., and Pastor, M. 2015. The haves, the have-nots, and the health of everyone: The relationship between social inequality and environmental quality. Annual Review of Public Health 36, 193–209. Farzin, Y.H., and Bond, C.A. 2006. Democracy and environmental quality. Journal of Development Economics 81, 213–35. Foster, A., and Kumar, N. 2011. Health effects of air quality regulations in Delhi, India. Atmospheric Environment 45, 1675–83. Garg, A. 2011. Pro-equity effects of ancillary benefits of climate change policies: A case study of human health impacts of outdoor air pollution in New Delhi. World Development 39, 1002–25. Grossman, G.M., and Krueger, A.B. 1995. Economic growth and the environment. Quarterly Journal of Economics 110(2), 353–77.

Holland, T.G., Peterson, G.D., and Gonzalez, A. 2010. A cross-national analysis of how economic inequality predicts biodiversity loss. Conservation Biology 23, 1304–13. Jorgenson, A., Schor, J., and Xiaorui, H. 2017. Income inequality and carbon emissions in the United States: A state-level analysis, 1997–2012. Ecological Economics 134, 40–48. Knight, K.W., Schor, J.B., and Jorgenson, A.K. 2017. Wealth inequality and carbon emissions in high-income countries. Social Currents 4, 403–12. McGee, J.A., and Greiner, P.T. 2018. Can reducing oncome inequality decouple economic growth from CO2 emissions? Socius 4, 1–11. Mehleb, R.I., Kallis, G., and Zografos, C. 2021. A discourse analysis of Yellow-Vest resistance against carbon taxes. Environmental Innovation and Societal Transitions 40, 382–94. Mikkelson, G.M., Gonzalez, A., and Peterson, G.D. 2007. Economic inequality predicts biodiversity loss. PLoS One, 5, e444. https://​doi​.org/​ 10​.1371/​journal​.pone​.0000444 Mohai, P., and Saha, R. 2015. Which came first, people or pollution? A review of theory and evidence from longitudinal environmental justice studies. Environmental Research Letters 10, 125011. Morello-Frosch, R., and Jesdale, B.M. 2006. Separate and unequal: Residential segregation and estimated cancer risks associated with ambient air toxics in U.S. metropolitan areas. Environmental Health Perspectives 114, 368–93. Okello, G., Devereux, G., and Semple, S. 2018. Women and girls in resource poor countries experience much greater exposure to household air pollutants than men: Results from Uganda and Ethiopia. Environment International 119, 429–37. Rubin, A.J. 2018. Macron inspects damage after ‘yellow vests’ protests as France weighs state of emergency. New York Times, December 1. Summers, L. 1992. Let them eat pollution. The Economist 322(7745), 66. Torras, M., and Boyce, J.K. 1998. Income, inequality, and pollution: A reassessment of the environmental Kuznets curve. Ecological Economics 25, 147–60. Wang, F., Yang, J., Shackman, J., and Liu, X. 2021. Impact of income inequality on urban air quality: A game theoretical and empirical study in China. International Journal of Environmental Research and Public Health 18, 8546. World Health Organization (WHO). 2021. Household air pollution and health. Fact sheet, 21 September. https://​www​.who​.int/​ news​-room/​fact​-sheets/​detail/​household​-air​ -pollution​-and​-health

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Industrial air toxics exposure in U.S. cities. Ecological Economics 107, 494–509.

34. Environmental ethics Environmental ethics simultaneously refers to both the inquiry into the right reasons to relate to non-human nature and the morality and specific values that inform environmental activism. Although precedents can be traced back through history to different worldviews and philosophies, environmental ethics essentially emerged as a critique of industrialisation and modern reification of nature during the 19th century, and boomed during the 1970s alongside the ecology crisis. As an academic discipline, environmental ethics has taken on the additional task of renewing ethical systems (Gosseries, 1998: 395) through the extension of the moral language—value, right, duty, responsibility, justice—to incorporate relations of non-reciprocity, specifically towards non-human nature and “distant others”. Together, the different schools of thought and theories can be summed up in three contiguous yet overlapping moments around the notions of intrinsic value— value in nature; ecological ethics—value of nature—; and climate ethics—responsibility in the face of global warming.

Intrinsic value in nature

The discipline of environmental ethics was conceived with the explicit objective of overcoming anthropocentrism, the meta-ideology underlying industrial society, where non-human nature is considered a mere means, devoid of value or self-interest, to serve human purposes (Eckersley, 1992: 3). From its very beginning, environmental ethics committed to a relational conception of nature, inspired by ecological science, where natural beings are conceived as part of the network or web of life, biotic community, or “biospherical net or field of intrinsic relations” (Leopold, 1949: 283; Naess, 1973: 95). Openly opposing methodological atomism and human–nature dualism, this holistic view informs the extension of the community of moral consideration to include non-human beings. The primary strategy in environmental ethics has been to deem that nature or at least some part of nature possesses intrinsic value, and that the environment is therefore not simply a medium to serve human interests. The notion of intrinsic value, a value that does not stem from or serve as the means

of another value, is not new. In moral philosophy, happiness, love, health, and human dignity are commonly considered ends in themselves, intrinsic (non-derivative) values. Similarly, the environmental ethicist argues that nature is more than an instrument and that it deserves moral consideration; therefore, nature generates for human beings duties towards non-human beings. In this regard, environmental ethics is facing at least three challenges: the identification of beings with intrinsic-natural value; the determination as to whether the domain of natural value coincides with that of environmental ethics; and the conclusion as to whether the latter can effectively overcome anthropocentrism. Firstly, different loci of natural value have been proposed. For ecocentrism, value is placed on totalities, such as ecosystems, the biosphere as a whole, or the Earth, including both its living parts and its non-organic components. For biocentrism, it is the specific life-forms that are valued. For zoocentrism or physiocentrism, value is attributed to non-human animals or at least those with sentience, the capacity to have a subjective experience of suffering. The reasons presented to justify the attribution of natural value differ in each approach: ecosystem integrity; to fulfil organisms’ potential; to have a good of its own; and to have an interest, among others (for a general overview, see Attfield, 2018; Jamieson, 1991; Callicott & Frodeman, 2009; Curry, 2011). Taken together, they are not necessarily compatible with each other. On the one hand, this is due to a conflict of interest among the different values: a given species (invasive, for example) can endanger the stability of an ecosystem or a landscape; the value of a species (endangered, for example) does not need to coincide with the aggregate value of the interests of its individual members in not suffering, but rather with its role within the ecosystem; and so on (Gosseries & Meijers, 2022). On the other hand, these cases of conflict present the problem of whether it is legitimate for humans to intervene to “protect” and “restore” natural values, both when the conflict has been induced by human action and when it takes place in environments with a low level of anthropisation or none at all. Therefore, any minimally practicable approach to environmental ethics is thus forced to establish some form of hierarchy for

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the different natural values, in other words, an axiological gradualism, and to ascribe more value to certain beings over others (Gosseries, 1998: 403). Secondly, insofar as environmental ethics goes hand in hand with an ecological conception of nature, the defence of the value of nature does not always form part of environmental ethics. Such is the case of the most widespread form of physiocentrism: sentientism. Its most widely influential version, inspired by Singer (1993), holds that animals with sentience are intrinsically valuable. Hence, the interests of non-human sentient animals must be treated with the same concern and respect to avoid the trap of human chauvinism—speciesism. From this perspective, the only site of natural value resides in the interest of specific individuals of human and non-human animal species in avoiding suffering. From this approach, a given species has no value other than that of the aggregate interests of its members in not suffering. That interest may be greater or lesser, depending on their sentience, and tends to be zero for most known species and natural entities. Therefore, other types of entities, such as the vegetal kingdom, and wholes, including species, biodiversity, ecosystems, or the natural environment, lack intrinsic value. The disagreements between the growing animal rights movement and environmentalism illustrate this tension, making it essential to define the boundaries between the sphere of environmental ethics and that of natural value. Thirdly, as we can see, the different approaches to environmental ethics must assume some form of axiological hierarchy. In most approaches, humanity (or part of it) is located at the top, under the assumption that its faculties are unique or greater than those of non-humans (self-awareness, sensitivity, rationality, morality, etc.), followed by other natural values that are ranked below it. In all the approaches, including those that most radically avoid human chauvinism—such as deep ecology—it is ultimately humans who identify the natural values. When all is said and done, environmental ethics attempts to balance the tension between overcoming anthropocentrism in the normative sense, as it does not reduce nature to its instrumental value for human beings, and anthropocentrism itself in the epistemic sense, as the valuators thereof are humans. This tension Joaquín Valdivielso

lies at the heart of the concept of intrinsic value. On the one hand, it assumes the a priori existence of natural values regardless of the valuator, which human beings recognise but do not assign. On the other hand, it understands that moral valuation is non-derivative and is therefore unconditioned. This is a problematic stance, for if natural values are considered non-derivative; they are not only independent of instrumental interests but also of human nature. A terminology more consistent with contemporary natural science should leave room for natural values as appraised and ranked vis-à-vis human nature, its history and evolutionary background, including moral emotions such as empathy and the capacity for altruism (Gosseries, 1998: 402; Tomassello, 2019). The root of this tension resides in the widespread assigned equivalence of instrumental and economic value within the field of environmental ethics, resulting in the use of the concept of intrinsic value. From the standpoint of practicality, the language of intrinsic value has been successful, as it has translated the aspiration to respect natural otherness for non-economic and non-chrematistic reasons. Beyond the level of philosophical analysis, it has been “the most powerful basis for the concerns of environmentalists” (Page, 2001: 170) and it has been key in justifying all sorts of environmental protection policies. Moreover, the concept of intrinsic value has been assumed increasingly in law to justify the legal personality of certain natural beings—such as great apes, sentient species, rivers, and “Earth Mother” seen as Pachamama, among many others—in the so-called rights of nature movement.

Ecological ethics: the value of nature

The difficulties presented by the language of intrinsic value have not prevented its predominance in environmental ethics and its widespread use in environmental studies in general. Even pioneers of ecological economics such as Daly and Cobb (1990: 200 ff) have assumed the view of intrinsic value of the biosphere and non-human life, asserting that “living things, individually and collectively, deserve consideration in their own right”. Nevertheless, the predominant perspective in ecological economics and in the

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social sciences in general is anthropocentric, or at least “for practical purposes the inanimate world may be viewed as mere means”, namely, as “resources” (Daly & Cobb, 1990: 200). This preferred strategy is based on the notion of the (human) right to the environment in a specific sense, as a socio-ecological metabolism. Insofar as this perspective views nature as an ecosystem, it can be defined as “ecological ethics”, yet this term is used with different meanings in different cases (Curry, 2011; Naredo, 2006). That being said, in the long run, this approach aspires to protect non-human nature significantly, indirectly, through instrumental measures aimed at human good (Pearce et al., 1991). The socio-metabolic approach focuses on the way in which environmental resources and services are appropriated and distributed among human beings. Indicators such as the Ecological Footprint, the Human Appropriation of Net Primary Production, the Material Footprint, and so on, are systematically used to illustrate material-ecological exchange as a space of unequal distribution of the environment and as a process of net consumption of natural capital, which is unsustainable over time (Sachs & Santarius, 2007; Martínez-Alier, 2002; Wiedman et al., 2013). This type of footprint expresses deeply asymmetrical distributions in the extraction and use of the flow of materials and energy and the allocation of benefits and burdens at the domestic, global, and intergenerational levels. Though it is not the norm, the interspecific dimension—distribution among species—is increasingly being incorporated into socio-metabolic metrics (Wilson, 2016). Given that all these domains of distribution are socially organised relationships, they can be scrutinised in terms of (in)justice and values and principles of fair distribution (Dobson, 2003). From this perspective, even the different views as to what the natural values are and the different languages of valuation of nature ultimately relate to conflicts over the distribution of the environment (Martínez-Alier, 2002). The principles of justice adduced to define this distribution as unjust vary in each case, although they usually stem from the notion of human rights (Barry, 1999; Bell, 2002; Skillington, 2017). Firstly, access to the goods of life-supporting systems is a precondition of human existence, health, physical integ-

rity, and well-being. Secondly, ecological inequality infringes on the principle of equal opportunity (including the right to recreation, aesthetic enjoyment, scientific research, etc.). Finally, it also contravenes this principle in terms of the future generations, as it leaves them less equivalent natural capital than what was available to the generations before them. In cases where it is considered that environmental goods can be distributed on an unequal basis—non-critical natural resources—Rawls’s difference principle is often applied; specifically, the clause requiring that it be for the benefit of the most disadvantaged. Rawls was not an environmental ethicist at all: he believed that the distribution of natural resources was not a problem, that “well-ordered” liberal societies sustain themselves with little, and there are no duties of justice at the post-national level (Rawls, 1999: 117). Nevertheless, his principle of difference has been deemed “inflammatory material” (Sachs and Santarius, 2007: 129) for the critique of the global socio-environmental metabolic order, as it generates both ecological and socio-economic inequality to the detriment of the most disadvantaged. The ultimate decision of what shall be subject to equal or unequal distribution will depend on the criterion of sustainability—in other words, whether sustainability is strong (meaning low or zero substitutability) or weak (high substitutability), and the corresponding intergenerational discount rate. From an ethical perspective, the answer to this question is determined by the specific principles of justice that are applied. If it is assumed that there is a wide range of instrumental values in nature—sensual delight, aesthetic, identity-cultural meaning, pedagogic, meaning of life (Krebs, 1999; Rawls, 1993: 245)—to which it is believed future generations should be just as entitled as the current generations, it would be appropriate to preserve all those “features of non-human nature whose loss would be irreversible” (Dobson, 1998: 47; Dobson, 2003: 164). Future generations should even have the option to morally consider natural values as part of a good life (Hailwood, 2005). Based on the principle of equal opportunity, the protection of irreversible nature therefore serves as a strategy of indirect protection under the criterion of strong sustainability and low or zero substitutability for many environmental Joaquín Valdivielso

208  Elgar encyclopedia of ecological economics Table 34.1 Principles of climate justice Principle

What?

Emissions egalitarianism Mitigation Polluter pays Beneficiary pays Ability to pay Grandfathering

Who? All

Why?

How?

Common ownership of the

Convergence in an equal and

atmosphere

sustainable share

Compensate disadvantage

In proportion to emissions

Adaptation

Responsible for climate

Mitigation

change

Adaptation

Advantaged by past

Compensate illegitimate

Mitigation

emissions

advantage

Adaptation

Those above the

Duties are proportional to

Mitigation

threshold

(economic) power

Mitigation

All

In proportion to benefits In proportion to wealth

Entitlement derived from

Equally (relative to status

past emissions

quo)

Source: Drawn up by author, based on Roser and Seidel (2017).

goods, thus providing extensive protection of the environment. Though this strategy broadens the range of goods to be protected far beyond the list of environmental resources and services in the quantitative indicators, it appears to be restrictive in relation to the initial expectations of environmental ethics, namely overcoming anthropocentrism.

Climate ethics: responsibility in the face of global warming

With the turn of the century, climate change has come to the forefront, virtually monopolising the environmental ethics debate. The idea of “climate ethics” encompasses different moral dilemmas relating to global warming. This particularly applies to the term “climate justice”, which stems from the Paris Agreement at the 2015 United Nations Climate Change Conference and Article 3 of the United Nations Framework Convention on Climate Change (UNFCCC) of 1992 on the commitment to undertake action to protect the climate on the basis of “common but differentiated responsibilities and respective capabilities”. Climate ethics present unique features (for an overview, see Dryzek et al., 2011; Gardiner et al., 2010). It is based on a metric of the anthropogenic carbon footprint (a measurement of equivalent tonnes of greenhouse gas emissions). The goal of preservation, namely, “preventing dangerous anthropogenic interference with Earth’s climate system”, has been foregone in favour of mitigation and adaptation, as global warming is considered to some extent irreversible. There is a relative—though not unanimous—consensus on the need to aspire to the lowest possible temperature increase Joaquín Valdivielso

scenario. The ethical reasons are often translated into practice through the global climate governance regime. Thus, climate ethics primarily focus on the question of what principle(s) of justice is/are the most plausible. In other words, despite the remaining open (epistemic) debate as to the existence of anthropogenic climate change in itself and the (normative) problems of whether action must be taken and how (Dryzek et al., 2013; Parfit, 1983), the predominant approach was born out of ecological ethics: the functions of the climate system cannot be carried out by means of other forms of capital—meaning low substitutability— and there are obligations of non-reciprocity with distant others. The discussion on the principles of climate justice (Table 34.1) stems from two premises. On the one hand, we are currently living in a harm scenario, where historically accumulated emissions have generated different intersecting inequalities (Skillington, 2017; Roser & Seidel, 2017: 9): the historical perpetrators of the emissions—generally more industrialised and affluent populations, countries, and regions—have a proportionally greater responsibility; and climate-change damage more heavily affects the poorest regions, future generations, women, and many non-human species and their habitats. On the other hand, total emissions should be distributed below a limit or maximum sustainable threshold that precludes catastrophic climate change, according to the values of responsibility and capacity. The different principles can be distinguished by the manners in which they respond to the questions of what should be

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done, to and by whom, for what reasons, and how responsibility is assigned to the correct stakeholder. It must be noted that at different times these principles have been implemented to a certain degree by the global climate governance regime. For instance, the polluter pays principle initially shaped the UNFCCC; the Kyoto Protocol and the “cap and trade” carbon offset market; and grandfathering was adopted by most EU member states in the allocation of emission allowances, which were given as a windfall profit to the sectors that have historically polluted the most. Each principle addresses the different ethical dilemmas in diverse ways, raising questions as to whether responsibilities are inherited, whether there is responsibility in case of ignorance of the harmful emissions generated, how transitional justice is applied, and how the costs of transition, mitigation, and adaptation are shared, among others. Taken together, they are not necessarily consistent with each other. On a practical level, climate governance leads to a combination of principles, as is the case in environmental ethics in general. In this case, however, the strong link between ethics and politics reveals the overlapping of other dimensions of justice. The climate governance regime has resulted in a slippery slope to a unilateral carbon metric and methodological nationalism—countries are the subjects. Besides, it displays significant democratic shortcomings: the interests of certain countries, corporations, present generations, and human interests are overrepresented (Eckersley, 2012), which partially explains their failure to reduce emissions. In other words, the algorithm of principles is the result of forms of interaction, deliberation, and negotiation that also ought to comply with the principles of procedural justice, and the representation of distant others, future generations, and non-human nature (Shrader-Frechette, 2002; González-Ricoy & Gosseries, 2016). The transition from an approach centred on natural values to climate ethics has resulted in a lowering of the fundamental objectives of environmental ethics. In addition, the pervasiveness of human intervention on the environment through climate change has enhanced the thesis that there is a human-independent nature no longer exists (Latour, 2004; Hamilton et al., 2015). Still,

environmental ethics is defined by a particular premise: the end of human exceptionalism and the false belief that human beings have transcended their natural condition and the limits imposed by the Earth’s metabolism (Schaeffer, 2007). Joaquín Valdivielso

References

Attfield, R. (2018). Environmental Ethics. A Very Short Introduction. Oxford: Oxford University Press. Barry, B. (1999). “Sustainability and Intergenerational Justice”. In A. Dobson (ed.), Fairness and Futurity: Essays on Sustainability and Justice. Oxford: Oxford University Press, 93–117. Bell, D. (2002). “How can political liberals be environmentalists?”, Political Studies, 50(4), 703–24. Callicott, J. B. & Frodeman, R. (eds.) (2009). Encyclopedia of Environmental Ethics and Philosophy. New York: Macmillan. Curry, P. (2011). Ecological Ethics. An Introduction. Cambridge: Polity Press. Daly, H. & Cobb Jr., J. B. (1990). For the Common Good: Redirecting the Economy toward Community, the Environment, and a Sustainable Future. Boston: Beacon Press. Dobson, A. (1998). Justice and the Environment. Conceptions of Environmental Sustainability and Dimensions of Social Justice. Oxford: Oxford University Press. Dobson, A. (2003). Citizenship and the Environment. Oxford: Oxford University Press. Dryzek, J. S., Norgaard, R. B. & Schlosberg, D. (2011). The Oxford Handbook of Climate Change and Society. Oxford: Oxford University Press. Dryzek, J. S., Norgaard, R. B. & Schlosberg, D. (2013). Climate Challenged Society. Oxford: Oxford University Press. Eckersley, R. (1992). Environmentalism and Political Theory. Toward an Ecocentric Approach. Albany: State University of New York Press. Eckersley, R. (2012). “Moving forward in the climate negotiations: multilateralism or minilateralism?”, Global Environmental Politics, 12(2), 24–42. Gardiner, S. M., Caney, S., Jamieson, D. & Shue, H. (eds.) (2010). Climate Ethics. Essential Readings. Oxford: Oxford University Press. González-Ricoy, I. & Gosseries, A. (eds.) (2016). Institutions For Future Generations. Oxford: Oxford University Press. Gosseries, A. (1998). “L’éthique environnementale aujourd’hui”, Revue Philosophique de Louvain, 96(3), 395–426.

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210  Elgar encyclopedia of ecological economics Gosseries, A. & Meijers, T. (2022). “Animal population ethics”. In G. Arrhenius, K. Bykvist, T. Campbell & E. Finneron-Burns (eds.), The Oxford Handbook of Population Ethics. Oxford: Oxford University Press, 546–68. Hailwood, S. (2005). “Environmental citizenship as reasonable citizenship”. In A. Dobson & A. Valencia (eds.), Citizenship, Economy and Environment. London: Routledge, 39–54. Hamilton, C., Bonneuil, C. & Gemenne F. (eds.) (2015). The Anthropocene and the Global Environmental Crisis: Rethinking Modernity in a New Epoch. New York: Routledge. Jamieson, D. (ed.) (1991). A Companion to Environmental Philosophy. Oxford: Blackwell. Krebs, A. (1999). Ethics of Nature. New York: Walter de Gruyter. Latour, B. (2004) Politics of Nature. How to Bring the Sciences into Democracy. Cambridge, MA: Harvard University Press. Leopold, A. (1949). A Sand County Almanac and Sketches Here and There. New York: Oxford University Press. Martínez-Alier, J. (2002). The Environmentalism of the Poor. A Study of Ecological Conflicts and Valuation. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Naess, A. (1973). “The shallow and the deep, long-range ecology movements: A summary”, Inquiry, 16, 95–100. Naredo, J. M. (2006). “Bases sociopolíticas para una ética ecológica y solidaria”, Polis, 13. http://​journals​.openedition​.org/​polis/​5430 Page, E. (2001). “Environmental ethics”. In J. Barry & E. G. Frankland (eds.), International Encyclopedia of Environmental Politics. London: Routledge, 168–71.

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Parfit, D. (1983). “Energy policy and the further future: The identity problem”. In A. D. MacLean & P. Brown (eds.), Energy and the Future. Totowa, NJ: Rowman & Allanheld, 166–79. Pearce, D. (ed.) (1991). Blueprint 2: Greening the World Economy. London: Earthscan. Rawls, J. (1993). Political Liberalism. New York: Columbia University Press. Rawls, J. (1999). The Law of Peoples. With “The Idea of Public Reason Revisited”. Cambridge, MA: Harvard University Press. Roser, D. & Seidel, C. (2017). Climate Justice. An Introduction. Abingdon: Routledge. Sachs, W. & Santarius, T. (2007). Fair Future: Resource Conflicts, Security, and Global Justice. A Report of the Wuppertal Institute for Climate, Environment and Energy. London: Zed Books. Schaeffer, J.-M. (2007). La fin de l’exception humaine. Paris: Gallimard. Shrader-Frechette, K. (2002). Environmental Justice. Creating Equality, Reclaiming Democracy. Oxford: Oxford University Press. Singer, P. (1993). Practical Ethics. Cambridge: Cambridge University Press. Skillington, T. (2017). Climate Justice and Human Rights. New York: Palgrave MacMillan. Tomassello, M. (2019). Becoming Human. A Theory of Ontogeny. Cambridge, MA: Harvard University Press. Wiedmann, T. O., Schandl, H., Lenzenc, M., Moran, D., Suh, S., West, J. & Kanemoto, K. (2013). “The material footprint of nations”, PNAS, 112 (20), 6271–6. Wilson, E. O. (2016). Half-Earth. Our Planet’s Fight for life. New York: Liveright.

35. Environmental footprints Origin of environmental footprints

Significant changes to the Earth’s environment have been witnessed over the past decades, with consequences that are undesirable or even disastrous for humanity (Barnosky et al. 2012; Scheffer et al. 2009). This has led to a transition from the Holocene (11 700 years ago)—a geological epoch that spanned a long period of time—to the Anthropocene (late 18th century)—a new era in which human disturbance is greatly eroding the stability and resilience of the Earth system (Crutzen 2002; Steffen et al. 2011). In striving to preserve the planet as a sustainable place for living and as a source of human welfare in this challenging era, there is a great need for novel approaches to modeling anthropogenic effects that are the key to identifying the driving forces of contemporary environmental change. Ecological footprint analysis (EFA) was originally introduced and advocated by ecological economists to evaluate the effects of anthropogenic activities on urban sustainability (Rees 1992). It compiled, on an area basis, the inputs of biological resources and the outputs of carbon emissions (i.e., the ecological footprint) and compared this to the regenerative and assimilative capacity of urban ecosystems (i.e., the biocapacity), indicating whether or not the situation remains sustainable (Rees 1997). At the human– environment interface, six types of land use on which human disturbance is most likely to occur have been taken into account: cropland, grassland, fishing ground, woodland, built-up land, and carbon uptake land. These relate to six ecosystem services, respectively: plant-based food production, animal-based food production, fish-based food production, timber production, living space supply, and carbon sequestration. In view of the success in raising public awareness of environmental issues and in evoking effective policy actions, Wackernagel and Rees (1997) have implemented an extension to the methodological application of the EFA, particularly pinpointing the nation-wide economy. In the latest edition of the National Footprint Accounts (NFAs), the “ecological footprint” is defined as the

area of biologically productive space required to produce the resources consumed and to absorb the waste generated, considering the prevailing technology and resources management practices (Borucke et al. 2013). The biocapacity, which probably can be traced back to the attempts that quantify human carrying capacity (Cohen 1995; Ehrlich 1982), is conceived in such a way that it can provide a region-specific threshold value for the ecological footprint of a given population. The comparison of the ecological footprint and biocapacity makes it possible to contrast sustainable and unsustainable consumption or production in an explicit manner.

Development of environmental footprints

Despite the worldwide popularity gained since its inception, the EFA is found incapable of capturing all aspects of human disturbance to the biosphere (Goldfinger et al. 2014), let alone to the atmosphere and hydrosphere. For this reason, a growing number of footprint-style indicators have been developed to complement the EFA in different dimensions. Examples in the environmental domain include the water footprint (Hoekstra and Hung 2002), the energy footprint (Stöglehner 2003), the carbon footprint (Wiedmann and Minx 2008), the chemical footprint (Hitchcock et al. 2012), the phosphorus footprint (Wang et al. 2011), the biodiversity footprint (Lenzen et al. 2012), the nitrogen footprint (Leach et al. 2012), the land footprint (Weinzettel et al. 2013), the material footprint (Wiedmann et al. 2015), and the resource footprint (Huysman et al. 2014), among others. In keeping with the ongoing expansion of the footprint family (Fang et al. 2014; Galli et al. 2012), the underlying methodologies are also undergoing rapid development. In addition to the NFA, life cycle assessment (LCA; Weidema et al. 2008), input–output analysis (IOA) (Hertwich and Peters 2009), material flow analysis (MFA; Schoer et al. 2012), and emergy-based (Zhao et al. 2005) and exergy-based methods (Chen and Chen 2007) have proved useful in calculating different environmental footprints at scales ranging from single products, processes, organizations, industries, and nations, to the whole human economy. Each tool has advantages and disadvantages. A case in point is LCA,

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whose comprehensive scope is useful for avoiding problem shifting, such as from one phase of the life cycle to another, from one region to another, or from one environmental problem to another. However, LCA struggles to address broader sustainability issues due to low data availability and the uncertainty of its results. Recently emerging hybrid approaches that take advantage of both LCA and IOA have shown great potential for meso-level footprint accounting for which well-established methods are lacking. As a whole, the choices of appropriate methods are playing a central role in quantitative footprint studies. The footprint family concept has attracted considerable interest within and outside in the domain of ecological economics, as it offers the scientific community an opportunity to achieve simultaneous measurement of various environmental footprints with implications for trade-off issues (De Meester et al. 2011; Ridoutt et al. 2014). By structuring a specified footprint family based on LCA (De Benedetto and Klemeš 2009) or on multi-regional input–output models (method developed on the basis of IOA; Ewing et al. 2012; He et al. 2022), for instance, current accounts for selected footprints have been conceptually integrated into single unified frameworks that would allow for greater transparency and consistency. A further step toward policy-relevant research is to develop an integrated footprint family that is supposed to encompass the complexity of some highly heterogeneous environmental issues, such as climate change (carbon footprint), water use (water footprint), land use (land footprint), and material use (material footprint).

Debates on environmental footprints

Since the first emergence of the ecological footprint, the term “footprint” has stimulated scientific debates, representing important steps in the ongoing discourse on sustainability (Hoekstra and Wiedmann 2014; Kitzes et al. 2009). Of particular concern is what actually counts as a footprint. Efforts have been made to lay the foundation for a widely accepted footprint definition (Hammond 2007; Pelletier et al. 2014). At least five com-

Kai Fang

peting criteria are available in the literature for definition of footprints: ● Footprints are indicators that express their results in an area-based metric unit. Clearly the ecological footprint represents the most obvious example of this. Many other well-known indicators, like the carbon footprint and the water footprint, would fall outside the scope, by this definition. ● Footprints are indicators that are intended for easy communication of results to the general public, policy makers, or other decision-makers. Indicators that aim to inform scientists and engineers, for instance, would in that case not be considered as a footprint. An example of such indicators is the eutrophication potential. ● Footprints are indicators that include a supply chain or that even take a full life cycle perspective; that is, indicators that only look at impacts of a country or company, for instance, without including at least the upstream impacts, would be disqualified from the footprint family with this categorization. ● Footprints are indicators that apply to the macro, economy-wide scale, in contrast to, for instance, LCA, which studies one functional unit (e.g., kg) of a product. ● Footprints are indicators that have the word “footprint” in their name. This criterion is based on an accidental nomenclature, but many footprint users implicitly stick to. It excludes, for instance, those studies that account for water footprint but refer to this as virtual water or embodied water. It is necessary to recognize further that a single footprint indicator may differ from another that has the same name but a different logic. For instance, the difference between volumetric water footprint and scarcity-based water footprint has been a subject of intense debate (Berger and Finkbeiner 2013). This highlights the importance of categorization in systematizing the footprint family—a topic that has grown in interest in recent years (Čuček et al. 2012). Fang et al. (2016) propose a classification scheme that successively divides the footprint family at four levels: the footprint is composed of two categories: resource and emission, each of which is separated into two levels (inventory and impact).

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This classification algorithm determines the relationship of footprint family members and is the key to making sense of the footprint concept. In addition, while integrating different footprint results, by using weighting factors, into a single composite metric is appealing from a user’s perspective, the scientific robustness and certainty of such a step are disputable, as any weighting schemes inevitably involve subjective judgments and are therefore prone to a lot of uncertainty (Huppes et al. 2012). The core of this challenge is the trade-offs between aggregate and disaggregate measures of environmental footprints (Fang and Heijungs 2015a).

Relation to life cycle assessment

Discussions on the close connection between environmental footprints and LCA have spawned an enormous literature (Hoekstra 2015; Lenzen 2014). A large number of pilot studies have provided concrete evidence of how LCA frameworks, in particular life cycle impact assessment, allow various footprint indicators to be suited for environmental impact assessment (Castellani and Sala 2012; Huijbregts et al. 2008) and absolute environmental sustainability assessment (ESA; Bjoern et al. 2020; Li et al. 2021). Given that the footprint community has indeed learned and borrowed much from LCA knowledge, some LCA experts argue that all footprint accounts should be exclusively on an LCA basis (Ridoutt et al. 2015), in the hope that the much broader appeal of footprint indicators could, in turn, facilitate the diffusion of life cycle thinking, particularly for non-LCA experts (Ridoutt and Pfister 2013; Weidema et al. 2008). In that case, LCA is meant to replace or supersede all other footprinting methods to be the “gold standard” (Fang and Heijungs 2015b). However, the relationship between environmental footprints and LCA is not as simple as it may seem. While being similar in many key elements, environmental footprints differ from LCA in that they can be operationalized in contexts where there is no clear life cycle or even without an LCA. This can be exemplified by the case of the NFA calculations, where data gaps constitute a major challenge to the use of LCA for nation-wide economy (Hellweg and Milà i Canals 2014). Moreover, there are certain important types of issues for which environmental footprints are desirable

while an LCA is not or only partially suitable, as proved by the risk of double counting in large-scale LCA (Lenzen et al. 2007), and vice versa, some typical impact categories (e.g., ozone depletion, ionizing radiation) are out of the scope of the footprint family in its present form. These discrepancies observed suggest that environmental footprints may preferably have a different orientation from LCA.

Relation to planetary boundaries

With the latest scientific knowledge, the planetary boundaries framework (PBF) defines the global-scale safe operating space for humanity by determining the difference between the current and threshold values for several environmental issues (Rockström et al. 2009; Steffen et al. 2015). The PBF, in this sense, has much in common with the EFA because both are well suited for monitoring the extent to which humanity is approaching or exceeding biophysical limits. Yet, this does not mean that there is no difference between the PBF and the EFA. Their differing optimal scales present one example, in which the former is primarily limited to the global scale with the aim of supporting global sustainability goals and pathways, whereas the latter has a far wider range of applications and thus can be applied to ESA at sub-global scales (e.g., cities, nations). In contrast to the ecological footprint, many footprint indicators, especially those based on LCA, do not include a comparison to any reference conditions, although this is increasingly being perceived as essential for ESA (Hoekstra 2015; Moldan et al. 2012). For this reason, allocating planetary boundaries to the national- or regional-scale shares in comparison to environmental footprints may be a novel way to enhance the policy relevance of EFA (Fang et al. 2015; Zhang et al. 2022). One challenge is that not all environmental issues are likely to show explicit threshold behaviors at sub-global scales. Meanwhile, lots of well-grounded footprint models would allow the PBF to have more accurate and reliable estimates of environmental pressures or impacts due to human activities at multiple scales. All this calls for a deeper understanding of the synergies between environmental footprints and planetary boundaries, and this will potentially present a useful tool for monitoring the progress and gaps of the Kai Fang

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Sustainable Development Goals in support of local to global sustainability governance (Chen et al. 2021; Fang et al. 2018; Vanham et al. 2019; Wiedmann and Allen 2021)

Future lines of research

The entry aims at bringing clarity and transparency to a number of unresolved and important issues regarding environmental footprints; nevertheless, there remain many gaps in our knowledge. To proceed with the development of environmental footprints, we offer a research agenda that includes the following prioritized topics: ● Harmonization of the structure, terminology, and notation of environmental footprints; ● Identification of the applicability and limitations of each environmental footprint; ● Evaluation of footprinting methods on scientific quality, policy relevance, and public acceptance; ● Development of approaches for uncertainty and sensitivity analysis of footprint models; ● Development of systematic and dynamic frameworks of the footprint family to integrate other sustainability-related (e.g., social, economic) footprints; ● Exploration of the methodological synergies with LCA, IOA, energy-based methods (e.g., emergy, exergy), etc.; and ● Improvement in the consistency and compatibility of the footprint and boundary metrics.

Acknowledgments

This work was supported by the National Natural Science Foundation of China (72074193), the Key R&D Program of Zhejiang Province (2022C03154), and the Ecological Civilization Project of Zhejiang University. Kai Fang

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36. Environmental governance Introduction

In ecological economics, much of the research on environmental governance has been conducted by scholars that are interdisciplinary across the social sciences, rather than crossing the boundary between ecology and economics. They have been mostly European and variably influenced by social economics, new institutional economics, institutional law and economics, and critical political economy/political ecology (Røpke, 2005; Paavola & Røpke, 2015). They have sometimes self-identified as “institutional ecological economists” (Paavola & Adger, 2005) or “critical institutional economists” (Vatn, 2017). These ecological economists acknowledge the importance of institutions, power, plural values, transaction costs, and the need for learning and preference formation, for example, which undermine traditional efficiency-based evaluation of policies and projects as well as monetary valuation of environmental goods and bads (Paavola & Røpke, 2015). This entry suggests that environmental governance is best understood as the establishment, reaffirmation, or change of institutional arrangements to resolve conflicts over environmental resources, such as forests and fisheries, biodiversity, atmospheric sinks for greenhouse gases, environmental safety, and quality of environmental media. The next section explains why the choice of institutional arrangements is a matter of social justice rather than economic efficiency. The following two sections offer a deeper look at multilevel and polycentric governance as more complex forms of environmental governance that have emerged as responses to large-scale environmental challenges, such as loss of biodiversity, climate change, and land degradation and desertification, over the past several decades, forming a complex institutional framework of overlapping and interacting institutional arrangements. Environmental governance is a concept akin to “sustainable development”. It provides a degree of integration across various perspectives, interests, and approaches, and yet it means different things for different people (Lange et al., 2013). For some, gov-

ernance refers to new ways of achieving social objectives in which states may participate but do not play a leading role (e.g. Rhodes, 2017; Stoker, 1998). To others, environmental governance encompasses all attempts to address environmental challenges or to resolve environmental conflicts by creating, changing, or reaffirming institutional arrangements (see Davidson and Frickel, 2004; Paavola, 2007). Further understandings of environmental governance are also possible. However, whichever standpoint is adopted, one clear observation emerges. The institutional framework of environmental governance has both thickened and become more complex in the past several decades (Jordan, 1999; Mitchell, 2003), for example, in the sense that the processes through which policies are designed, delivered, and implemented now involve greater interaction between a wider range of actors operating at an increased number of levels. The governance literature often distinguishes between “governance” and “government” by considering the absence of a coercive state power as the hallmark of “governance”. Yet governance is what governments do (Paavola, 2007: 94). Sometimes – as when local resource users govern themselves under customary institutions – environmental governance does not involve the state. Yet the customary resource users perform the governmental functions of legislation, administration, and adjudication and the government is involved, as the term “self-government” conveys. Rather than a monolithic external actor, the government, or the state, should be understood as sets of important arenas and instruments of collective action that are pertinent for environmental governance. The key implication of the involvement of the government or the state is that it entails a different distribution of power than self-governance solutions (Lange et al., 2013). Otherwise, national environmental and natural resource use policies perform similar functions and rely on comparable institutional solutions to those of the customary common property arrangements despite being formal, having larger jurisdictions, and relying on the mandatory powers of the state for enforcement. In this entry environmental governance is understood as the resolution of conflicts of interest over environmental resources through the establishment, reaffirmation, or change of institutional arrangements (Paavola, 2007;

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Paavola and Adger, 2005; Vatn, 2016; for the conflict theory of institutional change, see Knight, 1992). As an analytical concept, “environmental conflict” does not here refer to the existence of strife or open conflict between two or more parties (for examples of that view, see Diehl and Gleditch, 2001; Scheidel et al., 2020): it refers to conflicts of interests, situations where different interests in the environment are incompatible and cannot be satisfied simultaneously (Schmid, 2004; see also Scheidel et al., 2018: 587). Therefore, a choice has to be made on which interests to affirm and which to frustrate, and to what degree, if their balancing is possible (Schmid, 2004; see also Scheidel et al., 2018: 587). That choice is effectuated by affirming, changing, or establishing institutional arrangements, with attendant consequences for the distribution of economic and environmental goods and bads. The latter have also been termed “ecological distribution conflict” by Martínez-Alier and O’Connor (1996; see also Scheidel et al., 2018), a complementary view that gives more emphasis on the inequality of (environmental) outcomes than on the related conflict of interest further upstream. Coordination is also sometimes suggested as the reason for the existence of institutions (Taylor, 1987; Vatn, 2005) and institutions do indeed help achieve coordination. But many coordination problems in fact involve an interest conflict as an important aspect. When several ways of conducting matters exist, and one of them has to be chosen, this choice typically entails differential costs and benefits to involved and affected actors. Under different coordination solutions the beneficiaries and losers are likely to differ. Therefore, an interest conflict arises over which coordination solution and attendant distribution of costs and benefits to realise. But more importantly, all conflicts do not boil down to coordination problems: some of them are fundamentally about distributive and procedural justice (Paavola and Adger, 2005). An example is the past addition of lead to petrol – its allowing or prohibiting squarely maps on a choice between the economic interests of the oil industry and the interest of the general public in public health and safety. These conflicts are the most important drivers of institutional change because they demand institutional responses that settle them one way or another (see Knight, 1992). This does Jouni Paavola

not mean that social justice is achieved, it means that institutional choices addressing interest conflict generate a certain distribution of economic and environmental goods and bads. The choice between two institutional alternatives is a choice over two such distributions where winners and losers are different. These choices are not about efficiency, they are normative decisions on how to strike a balance in ecological distribution conflicts, and ones affected by the distribution of transaction costs and other sources of power in the fora where institutional choices are made. For example, REDD+ and many conservation policies strive for desirable environmental outcomes but do not sufficiently consider and support viable long-term economic alternatives for local livelihoods, in a context where local communities are not well placed to protect their interests. Environmental conflicts take place over the use and conservation of environmental resources, which include conventional natural resources, such as fisheries and forests, but also biodiversity, the ozone layer, and the global atmospheric sinks for greenhouse gases. Environmental resources also encompass environmental safety and the quality of environmental media such as water and air (Paavola, 2007). Many environmental resources are multifunctional in the sense that they generate multiple flows of ecosystem services for multiple beneficiaries located at different scales: this means that different ecosystem services provided by the same resource may have different “benefit (or cost) catchments”, areas within which the benefits (burdens) of ecosystem service flows are realised and appropriated (Paavola, 2007; Paavola & Primmer, 2019). Environmental conflicts can thus emerge over an individual ecosystem service (burden) flow: for example, over which of the competing irrigators can divert water from a watercourse to a consumptive water use. But environmental conflicts can also emerge as a result of claims to different ecosystem service flows. For example, claims to consumptive use of water for irrigation and claims to recreational in-stream uses of water can be in conflict with each other, as can be discharge of pollutants and consumptive water use. An advantage of a conflict-based definition of environmental governance is that it is analytically more encompassing than descriptive definitions emphasising the absence of gov-

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ernment or its limited role as the hallmark of environmental governance. The former helps consider all formal and informal institutions from norms, conventions, and customary common property arrangements to formal private property rights and markets, national environmental policies and regulations, and multilateral environmental agreements and regimes, as potential instruments for resolving conflicts, without excluding self-governance and networks where the state is not a central actor (see Paavola, 2011). The approach can underpin institutional analysis that can add granularity to our understanding of institutional diversity. For example, Vatn (2015) has demonstrated how market governance can assume multiple and sometimes incomplete forms, and that some governance solutions, such as Payment for Ecosystem Services schemes which are frequently considered examples of markets, actually do not meet the conditions for being a market. Some governance institutions, such as customary common property arrangements, local zoning provisions, and land use planning processes, organise and operationalise governance functions and processes at a single spatial scale. Multilevel environmental governance (MLEG) institutions organise and operationalise governance functions and processes at several spatial scales. Other forms of environmental governance include polycentric, network, interactive, and adaptive governance (Partelow et al., 2020). In what follows, a deeper look is offered into multilevel and polycentric governance as increasingly adopted approaches based on complex institutional arrangements.

Multilevel governance

MLEG solutions can emerge as a result of either bottom-up or top-down processes. Bottom-up processes are typically based on voluntary collective action, and the early literature on common property arrangements demonstrated how they can give rise to federations or other overarching institutional arrangements to coordinate the functioning of smaller-scale governance solutions (Ostrom, 1998). Ostrom (1990) discussed informal federations of irrigator associations as examples of them. Comparable but more formal governance arrangements have emerged to coordinate local governance efforts in the context of fisheries and shell fisheries (Berkes, 1992;

Hanna, 1998). Blomquist (1992) in turn explored how formal multilevel governance arrangements for the management and use of groundwater aquifers emerged in California through bottom-up negotiation processes. Another strand of literature demonstrates how many formal multilevel governance solutions have been created by top-down legal processes (Mitchell, 2003). For example, the United Nations Framework Convention on Climate Change and other multilateral environmental agreements require national actions, programmes, or solutions for the planning, coordination, and implementation of internationally agreed actions (Paavola, 2016). There is little consensus on why these multilevel structures emerge. Realist international relations scholars consider that the nation states first became environmentally conscious, and then started acting collectively to resolve their shared environmental problems by adopting multilateral environmental agreements. World systems scholars in turn argue that political globalisation has preceded, and been the driver of, national environmental management (Longhofer et al., 2016). They and constructivists also suggest that environmental science has fostered international environmental governance by generating shared rationalisations of environmental problems (Meyer et al., 1997). The questions of how and why MLEG emerges are important, because answers to these questions may be linked to the design and nature of MLEG institutions they give rise to. Top-down processes often generate multilevel institutional structures where smaller jurisdictions are nested within a larger jurisdiction(s). Hooghe and Marks (2003) call these kinds of multilevel governance solutions based on permanent, general-purpose jurisdictions with few levels and non-intersecting membership “Type 1”. Examples of Type 1 solutions include the federal state and many environmental policies in federal political systems. Hooghe and Marks (2003) also identify “Type 2” multilevel governance solutions, which have non-permanent and special-purpose jurisdictions, numerous levels, and intersecting memberships. Special districts for the provision of public services are examples (see Blomquist, 1992; Foster, 1997). These kinds of governance solutions are likely to emerge through bottom-up negotiation processes. Jouni Paavola

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Paavola (2016) examined potential explanations for emergence of multilevel environmental governance, which include 1) collective action challenges, 2) governance costs, 3) path dependence in institutional dynamics, and 4) multifunctionality of environmental resources. The collective action explanation has it that multilevel solutions may be needed to sustain collective action in large groups of actors involved in or affected by the use of large environmental resources, while governance cost explanation has it that multilevel arrangements may minimise governance costs or make them more affordable (Paavola, 2016). Path dependence reasoning suggests that multilevel solutions just replicate the structuring already in existence in other contexts, while multifunctionality explanation considers that the multitude of benefit and cost streams generated by environmental resources and their use require complex institutional arrangements encompassing multiple spatial scales (Paavola, 2016). These explanations are not mutually exclusive, although some may have more explanatory power with regard to specific multilevel governance arrangements.

Polycentric governance

The distinction between Type 1 and Type 2 governance solutions proposed by Hooghe and Marks (2003) is linked to the notion of polycentricity proposed by Vincent Ostrom and his colleagues (see Ostrom et al., 1961; Ostrom, 1972), who sought to explain the complex metropolitan governance structures that had emerged in the post-war decades for public service delivery in the United States. These new structures did not have a single core, which had characterised conventional governmental arrangements. Ostrom and his colleagues sought to establish the rationale of this kind of polycentric order, which he defined as “one where many elements are capable of making mutual adjustments for ordering their relationships with one another within a general system of rules where each element acts with independence of other elements” (Ostrom, 1999: 57). Polycentric order may emerge in a bottom-up way when diverse actors around a phenomenon like climate change seek to realise diverse benefits (or to avoid diverse costs) that accrue on different scales (see Ostrom, 2010): for example, mitigation actions do not Jouni Paavola

only generate the global benefit of reduced greenhouse gas emissions and reduced rate of climate change, but they also create co-benefits such as better air quality, reduced reliance on fossil fuels, reduced exposure to their price fluctuations, and improved energy security. These benefits can be a sufficient motivation for mitigation actions, although perhaps not on a comprehensive scale. The key interest of Vincent Ostrom was the horizontal dispersion of authority to govern. It was at the time a novel phenomenon, one which the established notions of government and governance were not well placed to account for. In light of the model proposed by Hooghe and Marks (2003), and the arguments of Ostrom and others regarding polycentricity, there is a continuum of horizontal dispersion of authority from monocentric to polycentric solutions, with hybrid solutions lying somewhere in the middle (Lemos and Agrawal, 2006). Hybrid forms of governance are created by many international environmental conventions: they are explicitly constituted as special-purpose jurisdictions vested with limited decision-making and other powers, but they often rely on national and sub-national general jurisdictions at lower levels of governance. But vertical structuring of governance is also involved in the examples Ostrom et al. (1961) and Ostrom (1972) discuss; it could be considered to range from full vertical symmetry to vertical differentiation. Early examples of polycentric governance that Vincent Ostrom and his collaborators were studying included voluntary cooperation of local governments to provide public services like public transport and waste management. Today, bottom-up and top-down processes are generating in many areas of public policy a mosaic of institutional diversity that includes state-based, hybrid, and voluntary measures that operate at levels from local to international and across levels (see Paavola, 2011). Climate change governance is perhaps the most prominent but not the only example. The international cornerstones of climate change governance play a role and are covering more greenhouse gas sources, include more ambitious emission-reduction targets, and address adaptation and its financing. National climate change policies implement international agreements and pursue domestic goals. Insurance and risk-sharing arrangements for adaptation are likely to

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demand public–private cooperation and to be based on hybrid solutions. Public–private cooperation and hybrid solutions are also likely to underpin mitigation-focused activities, particularly those related to carbon markets and experimental technologies such as carbon capture and storage and hydrogen. Regional and local governments are also increasingly be involved in the delivery of mitigation and adaptation through planning, regulation, investment, and public service provision.

Conclusions

Environmental governance is best understood as the establishment, affirmation, or change of institutions to resolve environmental conflicts over the use and conservation of environmental resources. The choice of governance solutions is a matter of social justice rather than of economic efficiency because different institutional arrangements distribute costs and benefits differently and also position actors differently with regard to decision-making. Multilevel and polycentric environmental governance have emerged as a response to large-scale environmental challenges over the past several decades, forming a complex institutional framework of overlapping and interacting institutional arrangements. Jouni Paavola

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1. Oxford University Press: Oxford and New York. Røpke I. 2005. Trends in the development of ecological economics from the late 1980s to the early 2000s. Ecological Economics 55: 262–90. Scheidel A, Del Bene D, Liu J, Navas G, Mingorría S, Demaria F, Avila S, Roy B, Ertör I, Temper L, Martínez-Alier J. 2020. Environmental conflicts and defenders: A global overview. Global Environmental Change 63: 102104. Scheidel A, Temper L, Demaria F, Martínez-Alier M. (2018) Ecological distribution conflicts as forces for sustainability: An overview and conceptual framework. Sustainability Science 13: 585–98. Schmid, AA. 2004. Conflict and Cooperation: Institutional and Behavioral Economics. Blackwell Publishing: Oxford and Malden. Stoker G. 1998. Governance as theory: Five propositions. International Social Science Journal 50(155): 17–28. Taylor M. 1987. The Possibility of Cooperation. Cambridge University Press: Cambridge. Vatn, A. 2005. Institutions and the Environment. Edward Elgar: Cheltenham, UK, and Northampton, MA. Vatn A. 2015. Markets in environmental governance: From theory to practice. Ecological Economics 117: 225–33. Vatn A. 2016. Environmental Governance: Institutions, Policies and Actions. Edward Elgar: Cheltenham, UK, and Northampton, MA. Vatn A. 2017. Critical institutional economics. In Routledge Handbook of Economics, Spash C (ed.). Routledge: London; 29–38.

37. Environmental input–output analysis 37.1 Introduction

The United Nations Conference on Environment and Development held in Rio de Janeiro in 1992 provided the fundamental principles and the programme of action for achieving sustainable development. It already identified unsustainable patterns of production and consumption as the major causes of the continued deterioration of the global environment and stated the need to change the way societies produce and consume to achieve a global sustainable development. Economic activities in its various stages have different effects on the environment that usually are considered externalities of the economic system since they are not reflected in market transactions. Nevertheless, these ‘environmental externalities’ exist precisely because the economic system is not isolated from the environment. These linkages, however, have seldom been considered in the past probably because the impacts of the economic activity were comparatively less harm

Source:

than now (Ayres and Kneese, 1969). But now the deterioration of the global environment has become an urgent and central issue, not only in politics and social spheres but also in academia. Generally, economists have considered the economy as a closed system in which producers’ and consumers’ decisions are coordinated by prices determined in the market. Although, with a high level of simplification of reality, the circular flow of income reflects this idea (Figure 37.1). In its simplest version, firms (producers) provide households (consumers) with goods and services in exchange for consumer expenditure, and households provide firms with factors of production in exchange for payment. Thus, this model describes the physical and the associated monetary flow between households and firms. This simple model can include not only the financial, government, and foreign sector, but also the linkages between the economy and the environment. Figure 37.2 shows how the economy is part of a broader context, the ecosystem or environment, in which solar energy enters allowing for different biogeochemical cycles that are basic requirements for the continuity of natural life. It is in this context where economic activity relates to the environment.

Author’s elaboration.

Figure 37.1 Circular flow of income of a two-sector economy

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On one hand, firms use not only labour and capital but also natural resources (that can be either renewable or non-renewable) to produce goods and services. On the other hand, production and consumption processes generate some waste that is given back in the environment. Part of this waste can be recycled within the economic system and reused again as a factor of production; however, in industrial economies, most of this waste is placed in the environment, which is one of the major causes of the deterioration of the environment. Moreover, through economic growth, the economic subsystem (grey box) may expand its physical dimension and may assimilate into itself a larger and larger proportion of the ecosystem (dotted grey area). This reflects that the economy usually needs more space, whether it be living space for an expanding population or space taken over to provide raw materials and sinks for waste

Source:

(Daly, 1999, p. 635). However, the economic growth may also be compatible with environmental improvement to some extent since, for instance, new and less polluting technologies can be introduced. Figure 37.2 places the economy and the environment within the same framework, but how can the monetary world (the economy) and the physical world (the environment) be analysed together within the economic discipline? Input–output analysis offers a suitable approach to study not only the interdependences inside the economy but also the interrelationships between the economy and the environment. Leontief (1928) pointed out that this approach to examining this kind of interaction opens the way for studies that deal with economic production and with other aspects, such as the effects of production and consumption on the environment. This capac-

Author’s elaboration from Daly (1999).

Figure 37.2 Physical flows between the economy and the environment

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Environmental input–output analysis  225

ity is due to three main features: first, the capacity of input–output analysis for combining monetary and physical units; second, input–output analysis allows for incorporating different assumptions that are more compatible with environmental policy objectives; and finally, the versatility of input–output models to analyse environmental problems at different levels (i.e. a single firm, economic sectors, a whole country, or the world) within the same approach.

37.2

The first attempts of environmental input–output models

In the late 1920s and 1930s Wassily Leontief established the foundations of input–output analysis,1 which have been continuously enriched not only by Leontief’s work but also by numerous contributions of other economists. Input–output analysis focuses on mutual interrelations between various parts of the economic system, which are usually called interdependences. In fact, this idea had already been developed earlier by Fran François Quesnay and León Walras;2 however, Leontief was the first to implement the general equilibrium theory (or general interdependence, as he preferred to term it) empirically and to adapt it to the needs of practical economic calculus. The hallmark of Leontief’s research was his continuous emphasis on the complementarity between theoretical reasoning and empirical work, being both two integrated parts rather than two separate disciplines within economics (Leontief, 1954, 1958, 1971, 1966/1986, 1987). Input–output analysis aims at bringing together the general laws and the observable facts of the economy. Leontief’s scientific methodology is summarised in his seminal papers (1928, 1936, 1937). In the 1960s and 1970s some economists tried to extend the standard input– output model to consider some of the links between the economy and the environment (Cumberland, 1966; Daly, 1968; Isard, 1969; Leontief, 1970a; Victor, 1972).3 Cumberland (1966) was the first to publish a theoretical input–output table that incorporated the economic and environmental interrelations. He tried to assign a monetary value for environment repercussions of economy activity to incorporate cost–benefit analysis

into the input–output model. However, he did not consider material flows between the economy and environment. Daly (1968) and Isard (1969) were the most ambitious authors since they tried to analyse bidirectional interactions between the economy and the environment and the interactions within the environment itself. Both designed a kind of input–output table composed of a matrix divided into four quadrants: the upper-left quadrant recorded the economic relations; the upper-right quadrant gathered the relationships from the economy to the environment; the lower-left quadrant focused on the relationships from the environment to the economy; finally, the lower-right quadrant compiled the relationships within the ecosystem itself. However, empirical applications of these models can hardly be done because of the difficulty of compiling data regarding the last quadrant.4 The difference between both models essentially lies in the structure of the table. Whereas Daly’s model followed the standard input–output table in which each sector produces only one product, Isard proposed a product-by-sector table that allowed for compiling secondary production of environmental output such as pollution. In 1970 Leontief published his own proposal to extend the input–output model to environmental pressures (Leontief, 1970a) and in 1972 he presented some results for the United States (Leontief and Ford, 1972). Leontief’s model only reflects the link from economy to environment, ignoring material flows in the reverse direction and within the environment. However, his simple proposal allows for estimating price effects produced by changes in anti-pollution technology and for analysing effects of government policies aimed at regulating industrial pollution. For doing so, Leontief introduced an extra row (indicating the amount of pollution generated by each sector) and an extra column (representing the abatement sector engaged in reducing the level of pollution) in the standard input–output table. The output of the new sector is the total amount of eliminated pollutant, and the corresponding final demand is the pollution tolerated by the economy.5 The last contribution was Victor (1972) who, due to the difficulty of gathering data about environmental interrelationships, only considered material flows from the environment to the economy and from the economy Mònica Serrano

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to the environment, discarding the lower-right quadrant in Isard’s model. The proposal by Victor (1972) is a conventional supply and use table framework extended to the environment that can be implemented without great difficulty. The system of National Accounting Matrix including Environmental Accounts (Keuning and Steenge, 1999) can be considered to some extent the heir of Isard’s and Victor’s approaches.

37.3

The input–output framework for a closed economy

According to Leontief’s scientific methodology, any empirical analysis needs to be based on a theoretical model but, at the same time, the theoretical terms of the model need to be directly observable. This section presents first the standard input–output table and its environmental extension that allows for computing the parameters of the input– output model, including the environmental extension, which is presented afterward. In this section we consider a closed economy because the multiregional input–output approach is the proper methodology to introduce exports and imports into the analysis. 37.3.1 The standard input–output table and the environmental extension The input–output table (Figure 37.3) describes the interconnections among sectors of an economy for any given period. Each

Source:

row describes the amount of a sector’s output that is distributed to other sectors and to final users; each column indicates the amount of input, both intermediate and primary, required to produce the total output of the corresponding sector. Figure 37.3(a) describes the standard input–output table where: ● Z ​ ​ is the intersectoral transaction matrix or intermediate consumption matrix that shows transactions among sectors, where​ Z​ij ​​epresents the flow of product from sector ​i​to sector ​j​. ● ​Y​shows the flow of products from each sector to final users, such as households (private consumption), government (public consumption), and investment (fixed capital formation). Often, all these categories are aggregated in one column vector called final demand (​ y)​.​ ● Primary inputs are those inputs that are not produced by sectors but are needed in the production process, such as labour, capital, and natural resources (e.g. mineral resources, land, water, etc.).6 Usually, natural resources are not considered together with labour and capital. In this sense, ​V​ would include only the primary input matrix of labour and capital whose element ​​vgi​  ​​represents the amount of input ​g​required to produce the output of sector ​j​. ● The lower-right quadrant would register primary inputs delivered directly to final

Author’s elaboration.

Figure 37.3 Input–output table: (a) Standard input–output table; (b) Environmentally extended input– output table

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Environmental input–output analysis  227

users (e.g. the value of compensation paid to household workers); however, in theoretical considerations, this quadrant is usually supposed to be zero without loss of generality. ● ​x​is the column vector of total output. That is, let ​i​be a column vector of ones; the column vector ​ x​is represented by​ x  =  Zi + y​. ● Finally, ​f​is the column vector that represents the total of primary inputs used in the economy (i.e. f​   =  Vi​). Although the transactions registered in the table are usually expressed in monetary values, originally the input–output table was conceived to register them in physical terms. In this case each sector and primary input should be measured in its appropriate unit. Then, steel can be measured in tonnes of standard product, electricity in kWh, computers in numbers of computers of average capability, labour in person-year, and so on. Although some service sectors can measure its output in physical units, it may be more useful to measure it in monetary values such as euros’ worth of output. When the input–output table is expressed in monetary units, the implicit physical unit is called the ‘Leontief unit’, defined as the quantity of product that we can buy for €1 in this economy.7 The fact that each commodity can be measured in its natural unit allows us to easily include and analyse the economy– environment interaction in the input–output framework, resulting in a variety of alternatives. When each sector of the input–output table is measured in its appropriate unit, it is called a mixed-unit input–output table; one kind of mixed-unit input–output table is the hybrid input–output table in which some rows, usually energy sectors, are measured in physical units, while the rest of the sectors are measured in monetary units. When all entries in the input–output table are measured in monetary values it is usually called a monetary input–output table. Finally, when the input–output table only registers product flows expressed in the same physical unit, such as tonnes, it is called a physical input–output table. Environmental input–output analyses can use hybrid and/or physical input–output tables, although more frequently they use monetary input–output

tables extended with environmental information in physical units,8 as Figure 37.3b describes. Here, Q ​ ​is a matrix of environmental ‘commodities’, either as an input flow from nature to the economic system (such as energy goods, water, mineral, or other material, etc.) or as output flow from the economic system to nature (i.e. air pollution, water pollutants, waste, etc.). Considering ​k​ different environmental ‘commodities’, the elements of this matrix ​qlj​  ​represent the amount of environmental ‘commodity’ ​l​generated by sector​ j​measured in its appropriate physical units. 37.3.2 The full basic input–output model and the environmental extension This input–output table allows us to apply directly the so-called full basic input–output models. The term ‘basic’ refers to the open, static input–output model proposed by Leontief (1944, 1946a, 1946b) in which final demand and factor prices are considered exogenous variables and in which changes over time are not considered endogenously in the model.9 The term ‘full’ is used to emphasise the fact that the input–output model consists of a quantity model, a price model, and an income identity (Duchin, 2004; Duchin and Steenge, 2007). Although input–output books usually present the price model as an addition of the basic input–output model (see Bulmer-Thomas, 1982; Miller and Blair, 2009), the increasing interest of environmental studies in applying input–output analysis brings us to reconsider this separation. When in the input–output table the output of each sector is measured in its ‘own’ unit, the quantity model determines the amount of product that each sector needs to produce to meet the final demand, whereas the price model determines the unit price of each product. If the input–output table is expressed completely in monetary values, the implicit physical unit is the so-called ‘Leontief unit’ (i.e. the quantity of product that can be bought by one monetary unit, such as euros), and consequently, the prices of all products that result from the price model are 1. However, even under these circumstances, the price model provides useful information about impacts on unit prices of technological changes, variations of factor prices, or introduction of environmental taxes. The latter should be of interest for some environmental studies that Mònica Serrano

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specifically incorporate environmental information measured in physical units. Formally, the full basic input–output model is defined by a system of five equations: ​x  ​=  (​ I ​−  A)​ ​− 1​ y

​(37.1a)​

​f  ​=  Lx

​(37.1b)​

​𝛍′ ​   ​=  ​𝛑′ ​L  

​(37.2a)​

​p′ ​   ​=  μ′​(​ ​I ​−  A​)​ ​− 1​ 

​(37.2b)​

​p′ ​y    ​=  ​𝛍′ ​x  

​(37.3)​

​r  ​=  Wx

​(37.4)​

Equation 37.1a, x​  =  (​ ​I − A​)​ −1​ y​, where​ I​is the identity matrix and ​​(​I − A​)​ −1​​is the Leontief inverse,11 tells us the amount of outputs ​x​required in this economy to meet a given final demand ​y​. Thus, x​ ​in Equation 37.1b, ​ f − Lx​ , indicates the amount of primary input (labour and capital) required to produce that particular set of output x​ ​. The Leontief inverse (​ ​I − A​)​ −1​ is the most useful and powerful tool in input–output analysis since it can reveal indirect effects within B​ the economy. Often denoted by matrix ​ element ​bij​ ​ represents the total output needed directly and indirectly from sector ​i​to satisfy one unit of final demand of sector j​​. In other words, each element ​bij​ ​ translates final demand for any product ​j​into required output from a sector ​i​. For instance, if final demand of sector ​j​increases by one unit, the initial effect indicates that sector j​ ​must produce one unit; however, to produce this unit, sector j​​ will demand inputs from a second sector ​i​ (i.e. the direct effect of sector j​​on sector i​​); then, to meet this new demand from sector j​​, sector ​i​will increase its production and will demand inputs from a third sector r​ ​ (indirect effect of sector ​j​on sector ​r​via sector ​i​), and so forth.

where the well-known basic input–output model (Equations 37.1a and 37.1b) is now renamed as the quantity input–output model; Equations 37.2a and 37.2b are the price input–output model; and Equation 37.3 is the income identity. Finally, Equation 37.4 is the environmental extension. In this system,​ A​ and ​L​are the parameters of the model, ​y​ 37.3.2.2 The price input–output model and the income identity and ​𝝅​the exogenous variables, and ​x​and ​p​ the endogenous variables determined by the Being ​𝛑′​a vector of cost (or quantity) of model. primary input per unit of output, we get vector ​𝛍′​from Equation 37.2a, ​​𝛍′ ​   =  𝛑′L​, whose elements represent the total monetary 37.3.2.1 The quantity input–output value (physical requirement) of all primary model 10 The matrix of direct input coefficients A ​ ​ and inputs required per unit of sector ​j​’s output, the matrix of primary input coefficients ​L​ which allows us to straightforwardly derive model are obtained directly from an input–output the solution of the price input–output −1 table by dividing all intermediate and primary (Equation 37.2b) ​​p′ ​   =  𝛍′​​(​I − A​)​  ​​(i.e. the inputs of each sector by the total output of this prices per physical unit of output). Product prices are proportional to the cost of primary sector. So, letting ​x​ ˆ​ be the diagonal matrix of inputs, and since ​L​is assumed to be constant, −1 −1 ​ˆ)​  ​  ​and ​L  =  V ​(​ x​ ​ˆ)​  ​  ​​, vector , we get ​A  =  Z ​(​ x​ Equation 37.2b also reveals that unit price of where the element ​aij​  ​(​ ​aij​  ​  =  ​zij​  ​ / ​xj​ ​)​​ represents each product results from the amounts of the the amount of output ​i​used as intermediate primary input ‘embodied’ in the production input to produce one unit of output ​j​; and of other products. That is, if labour were the element (​​ ​ ​lgj​  ​  =  ​vgj​  ​ / x​ j​ ​)​​is the amount of the only primary input, we could know how primary input g​ ​used as input per unit of sector​ much labour would be embodied in one unit j​’s output. The technology of this economy of product and, therefore, prices could be explained in terms of a ‘labour theory of can be designated by matrix ​T  =  [​ _A​L ​]​​, whose value’. column ​j​represents the average technology in Finally, we need to consider the income use of sector ​j​. or gross domestic product identity (Equation Mònica Serrano

Environmental input–output analysis  229

37.3) ​​p′ ​y    =  𝛍′x​, which is derived from the solutions of the quantity and price models, (i.e. 37.1a and 37.2b). This expression assures that the value of final demand ​p′y​equals total value added ​𝛍′x​.12 This identity should be satisfied not only in the base year for which the data have been compiled, but also under scenarios with different parameters and/or exogenous variables. 37.3.2.3 The environmental extension Finally, the total amount of environmental ‘commodity’ associated with a given vector of total output is represented by Equation , where ​ W​is the environ37.4, r​  =  Wx​ mental coefficient matrix whose element ​​wlj​  ​​ represents the environmental ‘commodity’ ​l​ emitted (or used) per unit of sector ​j​’s output, ​ˆ)​  ​ −1​​. From the and is defined by W ​   =  Q ​​(x​ demand perspective, Equation 37.4 can be expressed as ​r  =  W ​​(​1 − A​)​ −1​ y​or ​r  =  Jy​, where J​   =  W ​​(​I − A​)​ −1​is the matrix of total environmental intensity that shows direct and indirect environmental ‘commodities’ required to satisfy one unit of final demand of each sector. Mònica Serrano

Notes

1. Wassily Leontief won the Prize in Economic Sciences in Memory of Alfred Nobel in 1973 “for the development of the input–output method and for its application to important economic problems” (from the Nobel Foundation website: nobelprize.org). 2. Quesnay published the Tableau Economique in 1758 and Walras the Elements d’Economie Politique Pure in 1874. 3. The models mentioned are contributions within the input–output approach; however, in that time there was another contemporary work worth mentioning for its repercussions amongst economists. It was a paper published in the American Economic Review by Ayres and Kneese (1969), who attempted to adapt the Walras-Cassel general equilibrium model to include some of the environmental relations. Concretely, based on the physical law of conservation of mass, they showed that the production of residuals is an inherent and general part of the production and consumption process rather than a exceptional or a minor case that can be analysed as economic externalities. Although

they did not use an input–output approach, Ayres and Kneese contributed to opening the scope of the economy for considering the economy as an open system rather than a closed one. 4. Isard et al. (1968) tried to apply his model to the economy and environment of the Philadelphia Bay region. 5. Stone (1972) carried Leontief’s model a little further by introducing consumers explicitly in the model. 6. Capital goods (i.e. goods that are used for more than one single production period) are a singular case of ‘primary inputs’. In fact, the condition of ‘primary input’ can be questioned since they are produced by other industries; however, this interesting discussion is beyond the scope of this entry. 7. For instance, if the wage rate is €6 per hour of labour, the implicit Leontief unit of labour is 10 minutes, and the wage rate becomes €1 per 10 minutes of labour. 8. An example of a monetary input–output table that is environmentally extended is the NAMEA system. See Weisz and Duchin (2006) for a study on the differences between physical and monetary input–output tables in environmental input–output analysis. 9. The first version of the input–output model was the closed input–output model in which all the variables are endogenous (Leontief, 1937). It is a descriptive model that establishes the production and price structure of the economic system, but it does not define their respective levels. Between the closed and open input–output models there are the so-called partially closed input–output models, which internalise some components of the final demand. There are several ways do this; one is based on using a Social Accounting Matrix (SAM), which provides a much richer database than a simple input–output table. On the other hand, in contraposition to the static model, the dynamic input–output model introduces the investment decisions of each sector within the model (Leontief, 1970b). 10. It is also known as the technical coefficient, input– output coefficient, or intersectoral coefficient matrix. 11. It is also known as the multiplier matrix, total requirements matrix, or direct and indirect requirements matrix. 12. This identity should be known by those familiar with the linear programming representation of the full basic input–output model, in which the quantity model is the primary programme that maximises the total value added, and the price model is the dual programme that minimises the value of net final demand (the optimal solution of the dual programme will be the shadow price that should be equal to the opportunity cost of its inputs). A good reference for the input–output model and linear programming is Dorfman et al. (1958).

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of general interdependence”, The Review of Economic Statistics, 19(3), 109–32. Ayres, R.U., and Kneese, A.V. (1969) “Production, Leontief, W. (1944) “Output, employment, conconsumption, and externalities”, American sumption, and investment”, The Quarterly Economic Review, 59(3), 282–97. Journal of Economics, 58(2), 290–314. Bulmer-Thomas, V. (1982) Input–output analysis Leontief, W. (1946a) “Exports, imports, domesin developing countries: sources, methods and tic output, and employment”, The Quarterly applications, Chichester: John Wiley & Sons. Journal of Economics, 60(2), 171–93. Cumberland, J.H. (1966) “A regional interindustry Leontief, W. (1946b) “Wages, profit and prices”, model for analysis of development objectives”, The Quarterly Journal of Economics, 61(1), Papers in Regional Science, 17(1), 65–94. 26–39. Daly, H.E. (1968) “On economics as a life Leontief, W. (1954) “Mathematics in economics”, science”, Journal of Political Economy, 76(3), Bulletin of the American Mathematical Society, 392–406. 60(3), 215–33. Daly, H.E. (1999) “Steady-state economics: Leontief, W. (1958) “The state of economic avoiding uneconomic growth”, in J.C.J.M. van science”, The Review of Economics and den Bergh (ed.), Handbook of environmental Statistics, 40(2), 103–06. and resource economics, Cheltenham, UK, and Leontief, W. (1970a) “Environmental repercusNorthampton, MA: Edward Elgar Publishing, sions and the economic structure: an input– pp. 635–42. output approach”, The Review of Economics Dorfman, R., Samuelson, P.A., and Solow, R.M. and Statistics, 52(3), 262–71. (1958) Linear programming and economic Leontief, W. (1970b) “The dynamic inverse”, in analysis, New York: McGraw Hill. A.P. Carter and A. Bródy (eds.), Contributions Duchin, F. (2004) “Input–output economics and to input–output analysis: proceedings of the material flows”, Rensselaer Working Papers in fourth international conference on input– Economics n. 0424. output techniques, Amsterdam: North-Holland, Duchin, F., and Steenge, B. (2007) “Mathematical pp. 17–46. models in input–output economics”, Rensselaer Leontief, W. (1971) “Theoretical assumptions Working Papers in Economics n. 0703. http://​ and nonobserved facts”, American Economic www​.rpi​.edu/​dept/​economics Review, 61(1), 1–7. Isard, W. (1969) “Some notes on the linkage of Leontief, W. (1986 [1966]) Input–output economthe ecologic and economic systems”, Papers in ics (2nd ed.), New York: Oxford University Regional Science, 22(1), 85–96. Press. Isard, W., Bassett, K., Choguill, C., Furtado, J., Leontief, W. (1987) “Input–output analysis”, in Izumita, R., Kissin, J., Romanoff, E., Seyfarth, J. Eatwell, M. Milgate, and P. Newman (eds.), R., and Tatlock, R. (1968) “On the linkage of The New Palgrave. A Dictionary of Economics, socio-economic and ecologic systems”, Papers vol. 2, Macmillan: London, pp. 860–64. in Regional Science, 21(1), 79–99. Leontief, W., and Ford, D. (1972) “Air pollution Keuning, S.J., and Steenge, A.E. (1999) and the economic structure: empirical results “Introduction to the special issue on of input–output computations”, in A. Bródy ‘Environmental extensions of national accounts: and A.P. Carter (eds.), Input–output techniques, the NAMEA framework’”, Structural Change Amsterdam: North-Holland, pp. 9–30. and Economic Dynamics, 10(1), 1–13. Miller, R.E., and Blair, P.D. (2009) Input–output Leontief, W. (1928) “Die Wirtschaft als Kreislauf”, analysis: foundations and extensions (2nd ed.), Archiv für Sozialwissenschaft und Sozialpolitik, Cambridge: Cambridge University Press. 60, 577–623. English translation by R. Aylett, Stone, R. (1972) “The evaluation of pollution: “The economy as a circular flow”, Structural balancing gains and losses”, Minerva, 10(3), Change and Economic Dynamics, 1991, 2(1), 412–25. 181–212. Victor, P.A. (1972) Pollution: economy and enviLeontief, W. (1936) “Quantitative input and ronment, London: George Allen & Unwin. output relations in the economic systems of the Weisz, H., and Duchin, F. (2006) “Physical and United States”, The Review of Economics and monetary input–output analysis: what makes Statistics, 18(3), 105–25. the difference?”, Ecological Economics, 57(3), Leontief, W. (1937) “Interrelation of prices, 534–41. output, savings, and investment: a study in empirical application of the economic theory

Mònica Serrano

38. Environmental justice In 1982, residents of Warren County, United States, mobilised against the project of building a landfill for contaminated soils in their largely African-American community. This episode is considered to be the origin of environmental justice (EJ) movements, which are deeply rooted in Black history (350.org, 2023). The protests of Native American women against uranium mining affecting their reproductive health in the 1970s are also seen as forerunners of EJ activism, with gender playing a visible role in their dissent (Unger, 2004). The civil rights leader Benjamin Chavis would coin the term ‘environmental racism’ to refer to such cases of targeted exposure to environmental risks on the basis of race or ethnicity (Commission for Racial Justice, 1987). EJ emerged then as a type of environmentalism led by dispossessed people seeking social justice and equality in the access to a safe environment and equal protection for all communities (Agyeman, 2007; Pulido, 2017). In the following decades, new understandings of the nature of the dispossession would involve similar claims in new fields of contention (Figure 38.1). This chapter examines each of the stages in this expansion that has gradually enriched the vocabulary of EJ

scholarship and activism (Martinez-Alier et al., 2014). The first empirical EJ studies focussed on the issue of fair distribution of environmental protection and impacts of pollution. Since the late 1980s, several quantitative and spatial analyses confirmed a pattern of a disproportionate number of potentially toxic facilities (e.g., landfills, incinerators, industrial zones) located in areas with a majority of Black or Latino populations. Similar studies also demonstrated the exposure to pollutants of migrant agrarian workers (Agyeman et al., 2003; Bullard, 1994; Mohai and Saha, 2007). Assessing distributional issues is not without methodological controversies, one of them being the operationalisation of the equity vs equality gaps, placing the emphasis either on equal outcomes or on equal opportunities for all. Regardless of the challenges, persistent evidence of uneven distribution served as a lever for the political demands of the movements. In 1991, delegates from different grassroots movements adopted seventeen ‘Principles of Environmental Justice’ in Washington, D.C. (National People of Color Environmental Summit, 1991). With time, the US Environmental Protection Agency (EPA) would incorporate this political agenda into federal environmental programs, policies, and funding schemes (EPA, 2021). While distribution is still at the core of the EJ framework, the discussion soon articulated other inter-dependent dimensions (Schlosberg, 2013, 2007, 2004). Issues of

Figure 38.1 Environmental justice, an expanding framework

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recognition appear when the voice of those affected by the injustice is misrepresented or nullified, and they are therefore omitted from decision-making processes. By excluding cultural elements and specific identities that are critical to framing environmental values, expressions of difference and the complexity of human–environment relationships can be underplayed. This is common at the collective level in the case of Indigenous communities in matters that affect their cultural identity. In might also weaken the position of certain groups, such as women, children, older adults, or people with disabilities, in intra-community discussions. Technological and managerial solutions imposed from distant decision-making units may fail to recognise local difference due to ignorance or disrespect for biocultural diversity. Local knowledge systems, built on traditional ecological knowledge and site-specific socio-ecological settings, could suffer irreversible losses as a consequence. A way to address lack of representation is to demand direct participation in political or managerial decisions. This claim is frequent among community members or activists. Yet EJ movements can also rebel against participation procedures that embed lack of recognition or that do not provide participants with actual power (low levels of the ‘ladder of participation’; Arnstein, 1969). A critical approach to EJ equates participation to procedural justice, questioning its transformative influence without a deep understanding of power relations in EJ struggles (Svarstad and Benjaminsen, 2020). Indigenous peoples rarely participate directly in international institutions ruling international trade or in official discussions about the equitable share of benefits of natural resource use. Discrimination and linguicism are not the sole barriers for this participation. Acceptance of cross-cultural communication and learning is not easy in the prevalent systems of globalised consumption and production. When the aforementioned issues prevent communities from functioning in the way they want, a fundamental injustice occurs. People lose their capabilities to pursue their own imaginaries, and communities end up disappearing, symbolically and physically. This image is captured in Steve Lerner’s (2010) portrayal of ‘sacrifice zones’ in the US taking action in order not to be forgotten. The term, which refers to areas permanently Beatriz Rodríguez-Labajos

damaged by industrial development, has been taken up by environmental defenders against land grabs, mining, and industrial operations around the globe (Broto and Calvet, 2020; Oliveira and Hecht, 2016). According to the well-known capabilities approach, developed by Amartya Sen, and Martha Nussbaum, communities engaged in EJ struggles would be defending their ability to function within their own visions of well-being in face of such threats (Edwards et al., 2015). The dimensions just presented – distribution, recognition, participation, and capabilities – meant a theoretical and terminological expansion of EJ. This process lasted several decades during which EJ derived from individual-based to collective-based concerns and claims for justice (Figure 38.2). Along with the conceptual expansion, the EJ discourse spread worldwide, covered a wide range of themes, and gained popularity as a source of transformative politics (Schlosberg, 2013; Walker, 2009). Regarding the expansion of the geographic span, EJ became part of the demands of communities and organisations outside the US, especially in the Global South. Several cases reveal that international trade has transferred the material load of global economic growth between countries with different levels of development (Fan et al., 2020; Weber and Cabras, 2021). Meanwhile, Asia, Africa, and Latin America have increased their material consumption above the global average since the 1980s (Giljum et al., 2014). Increased domestic extraction tends to exacerbate existing distributional issues and recognition gaps within countries (from the uneven distribution of costs and benefits between local elites and communities in extraction areas) or between countries (from the deterioration in the terms of trade). The EJ discourse has broadened its thematic scope beyond the original struggles against pollution and waste. Conflicts over land grabbing and monoculture expansion (Busscher et al., 2020), or mineral extraction (Rodríguez-Labajos and Özkaynak, 2017) mobilise the argument of insufficient access to benefits by communities that bear instead the environmental and social costs of the projects’ operation. Reduced access to land and water are recurrent claims. In fact, unequal access to water at the desired quality among social groups has configured the fertile field of water justice (Boelens et al., 2018), in

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Source:

Expanded from the author’s contribution to Corbera et al. (2019).

Figure 38.2 Timeline of milestones in the history of environmental justice

which water is both medium and end goal of political power over socionatures. Similarly, unequal power relations over the global carbon cycle and the unequal distribution of costs of action and inaction against climate change frame the studies and actions about climate justice (Rodríguez-Labajos, 2013; Warlenius, 2017). Unsurprisingly, the notion of global EJ movements mushroomed the literature from the late 2000s onwards (Carruthers, 2008; Martinez-Alier et al., 2016; Sikor and Newell, 2014; Sze and London, 2008). There are different reasons to support this idea of a globalised movement transcending localised environmental struggles. First, networks

of activists operating in different geographies but related to the same types of material process may share concerns, strategies, and achievements. Aydin et al. (2017) make this case for mining conflicts around the world, demonstrating network effects in anti-mining struggles. Second, actors along commodity chains may connect resistances as well (e.g., when opponents to soybean monocultures in South America partner with critical consumers against industrial meat production in Europe). The telecoupling literature offers a framework to articulate such connections (Corbera et al., 2019). Third, the existence of a rich and shared vocabulary stemming from social movements across different types of Beatriz Rodríguez-Labajos

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conflicts (Martinez-Alier et al., 2014) speaks volumes about a shared political project as well. A sad manifestation of this decentralised power emerging from environmental conflicts is a global pattern of pressures put on environmental defenders via intimidation, sexual violence, criminalisation, and killings (Global Witness, 2021). Despite the expanding scope of the EJ discourse, the notions described above are still deeply rooted in the philosophical sources of the West. The political philosopher John B. Rawls (1958, 1971) established a reference frame for the distinction between justice and fairness. Justice involves societal practices and principles to overcome disadvantageous inequalities in achieving well-established rights. Closer to the idea of procedural justice that emerges from a social contract, ‘justice’ is a pact between self-interested parties to promote the common good. Clearly rooted in utilitarianism, Rawls himself saw it as not far from being ‘a kind of efficiency’ (1958: 184). Meanwhile, fairness refers to the moral context that enables such a pact. Fair choices satisfy every party’s conception of legitimate claims, even in the absence of ad hoc rules. The political liberalism embedded in these notions make them appropriate for debates in Western societies – less so for controversies that connect culturally diverse agents, and non-human ethics. The ontologies from First Nations or (traditional and current) Eastern views do not necessarily accommodate well with the standard EJ frame. Thus a distinctive Indigenous EJ approach is emerging in both scholarly literature (McGregor et al., 2020) and in public action (EPA, 2019), based on legal traditions and spiritual or religious ethics and the respect to traditional ecological knowledge. Recent calls for decolonising EJ in the Latin American contexts point towards the same direction (Álvarez and Coolsaet, 2020). The emergence of new political actors under the umbrella of EJ ended up reaching Mother Earth and all living beings as subjects of rights. The ongoing Anthropocene extinction may eliminate between half and three-quarters of Earth’s life forms. Thus, organisms threatened by extinction face a particular type of injustice due to human action (Wienhues, 2020). In this context, the notion of ecological justice (Baxter, 2005) applies concepts of distributional and procedural justice to non-human beings. The Beatriz Rodríguez-Labajos

parallels between this situation and issues of intergenerational justice linked to climate change would explain why grassroots movements such as Extinction Rebellion use the term ecological justice in their demands. EJ has succeeded in influencing public policies in the US and beyond. Clearly, multiple social movements around the world have adopted the EJ discourse. Data compilation on EJ conflicts globally depicts both a bleak picture of increasing number of EJ disputes and the hope of transformative capacity emerging from resistances (EJAtlas, 2023). Yet the EJ discourse is not free of several criticisms in terms of its real potential to reverse disruptive environmental pressures. The first one is the reactive nature of most EJ struggles, since communities tend to confront specific projects only after perceiving their threats. In this respect, a second criticism is the need of a fundamental critique of the mechanisms of dispossession that originated the injustice in the first place, which led some radical geographers to challenge the political nature of the EJ movements (Swyngedouw, 2009). Finally, the turn of EJ into the mainstream raises concerns of co-optation, that is, the possibility that EJ policies remain as cosmetic measures without respond to real concerns of communities and social movements (Harrison, 2015). Even worse, the EJ vocabulary and measures can be put at the service of greenwashing or the promotion of disruptive projects. A foremost example is the misuse of measures to promote prior informed consent in cases of potential environmental impact that are manipulated to validate, rather than hinder, locally unwanted projects. Aware of these challenges, EJ scholars and activists have progressively enhanced their work, bringing it to new topics and scopes of action. Today, the emergence of new EJ leaderships, especially that of younger generations, and the response to the challenge of an increasingly digitalised world, are two major lines of EJ development. As a final note, it is worth emphasising the strong connections between EJ and ecological economics. Clearly, the consideration of different languages of valuation underneath the well-known thesis of the environmentalism of the poor is closely connected with recognition issues that accompany distributive injustices. Ultimately, the social-metabolic origin of environmental injustices is a common topic of concern, which leads to a shared

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critique of economic growth as the basic driver of both environmental injustices and unsustainability. Beatriz Rodríguez-Labajos

of communities with hazardous waste sites. United Church of Christ, New York. Corbera, E., Busck-Lumholt, L.M., Mempel, F., Rodríguez-Labajos, B., 2019. Environmental justice in telecoupling research, in: Friis, C., Nielsen, J. (Eds.), Telecoupling. Palgrave Studies in Natural Resource Management, References pp. 213–32. Palgrave Macmillan, Cham. https://​ 350.org, 2023. The environmental justice movedoi​.org/​10​.1007/​978​-3​-030​-11105​-2​_11 ment is rooted in Black history. 350 Press Edwards, G.A.S., Reid, L., Hunter, C., 2015. Release. https://​350​.org/​black​-history​-month/​ Environmental justice, capabilities, and the Agyeman, J., 2007. Environmental justice and theorization of well-being. Progress in Human sustainability, in: G. Atkinson, S. Dietz, E. Geography 40, 754–69. https://​doi​.org/​10​.1177/​ Neumayer (Eds.), Handbook of Sustainable 0309132515620850 Development, pp.  171–88. Edward Elgar EJAtlas, 2023. Environmental Justice Atlas. Publishing, Cheltenham, UK, and Northampton, https://​ejatlas​.org/​ MA. Environmental Protection Agency (EPA), 2019. Agyeman, J., Bullard, R.D., Evans, B., 2003. Just Environmental justice for tribes and Indigenous Sustainabilities: Development in an Unequal peoples. https://​www​.epa​.gov/​e​nvironment​ World. MIT Press, Cambridge, MA. aljustice/​environmental​-justice​-tribes​-and​ Álvarez, L., Coolsaet, B., 2020. Decolonizing -indigenous​-peoples environmental justice studies: a Latin American EPA, 2021. Environmental justice. https://​www​ perspective. Capitalism Nature Socialism 31, .epa​.gov/​environmentaljustice 50–69. https://​doi​.org/​10​.1080/​10455752​.2018​ Fan, M.-F., Chiu, C.-M., Mabon, L., 2020. .1558272 Environmental justice and the politics of polArnstein, S.R., 1969. A ladder of citizen parlution: the case of the Formosa Ha Tinh Steel ticipation. Journal of the American Planning pollution incident in Vietnam. Environment Association 35, 216–24. and Planning E: Nature and Space 5, 189–206. Aydin, C.I., Özkaynak, B., Rodríguez-Labajos, B., https://​doi​.org/​10​.1177/​2514848620973164 Yenilmez, T., 2017. Network effects in envi- Giljum, S., Dittrich, M., Lieber, M., Lutter, S., ronmental justice struggles: an investigation of 2014. Global patterns of material flows and their conflicts between mining companies and civil socio-economic and environmental implicasociety organizations from a network perspections: a MFA study on all countries world-wide tive. PLoS One 12(7), e0180494. https://​ doi​ from 1980 to 2009. Resources 3, 319–39. .org/​10​.1371/​journal​.pone​.0180494 https://​doi​.org/​10​.3390/​resources3010319 Baxter, B., 2005. A Theory of Ecological Justice. Global Witness, 2021. Last line of defence. The Routledge, London. https://​doi​.org/​10​.4324/​ industries causing the climate crisis and attacks 9780203458495 against land and environmental defenders. Boelens, R., Perreault, T., Vos, J., 2018. Water Global Witness, London. Justice. Cambridge University Press, New Harrison, J.L., 2015. Coopted environmental York. justice? Activists’ roles in shaping EJ policy Broto, V.C., Calvet, M.S., 2020. Sacrifice zones implementation. Environmental Sociology 1, and the construction of urban energy landscapes 241–55. https://​doi​.org/​10​.1080/​23251042​ in Concepcion, Chile. Journal of Political .2015​.1084682 Ecology 27, 279–99. https://​doi​.org/​10​.2458/​ Lerner, S., 2010. Sacrifice Zones. The Front Lines V27I1​.23059 of Toxic Chemical Exposure in the United Bullard, R.D., 1994. Dumping in Dixie: Race, States. MIT Press, Cambridge, MA. Class and Environmental Quality. Westview Martinez-Alier, J., Anguelovski, I., Bond, P., Del Press, Colorado. Bene, D., Demaria, F., Gerber, J.-F., Greyl, Busscher, N., Parra, C., Vanclay, F., 2020. L., et al., 2014. Between activism and science: Environmental justice implications of land grassroots concepts for sustainability coined by grabbing for industrial agriculture and forenvironmental justice organizations. Journal of estry in Argentina. Journal of Environmental Political Ecology 21, 19–60. https://​doi​.org/​10​ Planning and Management 63, 500–22. https://​ .2458/​v21i1​.21124 doi​.org/​10​.1080/​09640568​.2019​.1595546 Martinez-Alier, J., Temper, L., Del Bene, D., Carruthers, D. V., 2008. The globalization of Scheidel, A., 2016. Is there a global environenvironmental justice: lessons from the U.S.– mental justice movement? Journal of Peasant Mexico Border. Society & Natural Resources Studies 43, 731–55. https://​doi​.org/​10​.1080/​ 21, 556–68. 03066150​.2016​.1141198 Commission for Racial Justice, 1987. Toxic wastes McGregor, D., Whitaker, S., Sritharan, M., 2020. and race in the United States. A national report Indigenous environmental justice and sustainon the racial and socio-economic characteristics

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236  Elgar encyclopedia of ecological economics ability. Current Opinion in Environmental Sustainability 43, 35–40. https://​doi​.org/​10​ .1016/​j​.cosust​.2020​.01​.007 Mohai, P., Saha, R., 2007. Racial inequality in the distribution of hazardous waste: a national-level reassessment. Social Problems 54, 343–70. https://​doi​.org/​10​.1525/​sp​.2007​.54​.3​.343 National People of Color Environmental Summit, 1991. The principles of environmental justice. http://​www​.ejnet​.org/​ej/​principles​.pdf Oliveira, G., Hecht, S., 2016. Sacred groves, sacrifice zones and soy production: globalization, intensification and neo-nature in South America. Journal of Peasant Studies 43, 251–85. https://​ doi​.org/​10​.1080/​03066150​.2016​.1146705 Pulido, L., 2017. Conversations in environmental justice: an interview with David Pellow. Capitalism Nature Socialism 28, 43–53. https://​ doi​.org/​10​.1080/​10455752​.2016​.1273963 Rawls, J., 1958. Justice as fairness. The Philosophical Review 67, 164–94. https://​doi​ .org/​10​.2307/​2182612 Rawls, J., 1971. A Theory of Justice. Harvard University Press, Cambridge, MA. Rodríguez-Labajos, B., 2013. Climate change, ecosystem services, and costs of action and inaction: scoping the interface. Wiley Interdisciplinary Reviews Climate Change 4, 555–73. https://​doi​.org/​10​.1002/​wcc​.247 Rodríguez-Labajos, B., Özkaynak, B., 2017. Environmental justice through the lens of mining conflicts. Geoforum 84, 245–250. https://​doi​.org/​10​.1016/​j​.geoforum​.2017​.06​ .021 Schlosberg, D., 2004. Reconceiving environmental justice: global movements and political theories. Environmental Politics 13, 517–40. https://​ doi​.org/​10​.1080/​0964401042000229025 Schlosberg, D., 2007. Defining Environmental Justice: Theories, Movements, and Nature. doi​ Oxford University Press, Oxford. https://​ .org/​10​.1093/​acprof:​oso/​9780199286294​.001​ .0001 Schlosberg, D., 2013. Theorising environmental justice: the expanding sphere of a discourse.

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Environmental Politics 22, 37–55. https://​ doi​ .org/​10​.1080/​09644016​.2013​.755387 Sikor, T., Newell, P., 2014. Globalizing environmental justice? Geoforum 54, 151–7. https://​doi​ .org/​10​.1016/​J​.GEOFORUM​.2014​.04​.009 Svarstad, H., Benjaminsen, T.A., 2020. Reading radical environmental justice through a political ecology lens. Geoforum 108, 1–11. https://​doi​ .org/​10​.1016/​j​.geoforum​.2019​.11​.007 Swyngedouw, E., 2009. The antinomies of the postpolitical city: In search of a democratic politics of environmental production. International Journal of Urban and Regional Research 33, 601–20. https://​doi​.org/​10​.1111/​j​.1468​-2427​ .2009​.00859​.x Sze, J., London, J.K., 2008. Environmental justice at the crossroads. Sociology. Compass. 2, 1331–54. https://​doi​.org/​10​.1111/​j​.1751​-9020​ .2008​.00131​.x Unger, N.C., 2004. Women, sexuality, and environmental justice in American history, in: Stein, R. (Ed.), New Perspectives on Environmental Justice: Gender, Sexuality, and Activism, 45–60. Rutgers University Press, New pp.  Brunswick, NJ. Walker, G., 2009. Globalizing environmental justice: The Geography and Politics of Frame Contextualization and Evolution. Global Social Policy 9, 355–82. https://​doi​.org/​10​.1177/​ 1468018109343640 Warlenius, R., 2017. Decolonizing the atmosphere: the climate justice movement on climate debt. Journal of Environment & Development 27, 131–55. https://​doi​.org/​10​.1177/​ 1070496517744593 Weber, G., Cabras, I., 2021. Environmental justice and just transition in the EU’s sustainability policies in third countries: the case of Colombia. The International Spectator 56, 119–37. https://​ doi​.org/​10​.1080/​03932729​.2021​.1946262 Wienhues, A., 2020. Ecological Justice and the Extinction Crisis. Giving Living Beings their Due. Bristol University Press, Bristol. https://​ doi​.org/​10​.47674/​9781529208528

39. The environmental Kuznets curve Introduction

The environmental Kuznets curve (EKC) is a hypothesized relationship between various indicators of environmental degradation and countries’ gross domestic product (GDP) per capita. In the early stages of a country’s economic development, environmental impacts and pollution increase, but beyond some level of GDP per capita (which will vary for different environmental impacts), economic growth leads to environmental improvement. This implies that environmental impacts or emissions per capita are an inverted U-shaped function of GDP per capita, whose parameters can be statistically estimated. Figure 39.1 shows a very early example of an EKC. A large number of studies have estimated such curves for a wide variety of environmental impacts ranging from threatened species to nitrogen fertilizers, though atmospheric pollutants such as sulfur dioxide and carbon dioxide have been investigated most. Panayotou (1993) was the first to call this relationship the EKC, where Kuznets refers

Note: Source:

to the similar relationship between income inequality and economic development proposed by Nobel Laureate Simon Kuznets, known as the Kuznets curve. Grossman and Krueger (1991) introduced the EKC in an analysis of the potential environmental effects of the North American Free Trade Agreement (NAFTA). They argued that the economic growth they expected to result from NAFTA would have beneficial effects on the environment. This contrasted with the views of some ecological economists, such as Daly (1993), who claimed that NAFTA would be very environmentally detrimental. The EKC also featured prominently in the 1992 World Development Report in which the World Bank (1992) endorsed the World Commission on Environment and Development (1987) view that the choice between development and the environment is a “false dichotomy” (p. 25) and that poverty reduction is essential for environmental protection. The EKC has since become the dominant approach among economists to modeling ambient pollution concentrations and aggregate emissions, has become very popular in policy circles, and is even found in introductory economics textbooks.

Environmental Kuznets curve estimated by Panayotou (1993). See Stern et al. (1996) for details.

Figure 39.1 An environmental Kuznets curve

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Critique

Despite this, the EKC was criticized almost from its beginning on empirical and policy grounds, and debate continues. It is undoubtedly true that some dimensions of environmental quality have improved in developed countries at the same time that they have become richer, while others have not. Problems that are more acute, visible, easier to mitigate, have fewer international and intergenerational spillovers, and have more impact on human health are more likely to be addressed. City air and rivers in developed countries have become cleaner since the mid-20th century, and in some countries forests have expanded. Emissions of some pollutants, such as sulfur dioxide, have clearly declined in most developed countries in recent decades. But there is more mixed evidence for pollutants such as carbon dioxide. Carbon emissions have fallen in the last 40 years in some developed countries, such as the United Kingdom or Sweden, while they have increased in others, such as Australia or Japan. There is also evidence that emerging countries take action to reduce severe pollution. For example, Japan cut sulfur dioxide emissions in the early 1970s following a rapid increase in pollution when its income was still below that of the developed countries (Stern, 2005), while China has also acted to reduce sulfur emissions and particulate pollution in recent years. As further studies were conducted and better data accumulated, many of the early econometric studies that supported the EKC were found to be statistically fragile. Initially, many understood the EKC to imply that the best way for developing countries to improve their environment was to get rich (e.g. Beckermann, 1992). This alarmed others (e.g. Arrow et al., 1995) as, while this might address some issues like deforestation or local air pollution, it would likely exacerbate other environmental problems such as climate change. Even if there is an EKC for per capita impacts, environmental impacts would increase for a very long time if either the majority of the population is on the rising part of the curve, population is growing, or both of these are true (Stern et al., 1996).

David I. Stern

Explanations

The existence of an EKC can be explained either in terms of deep determinants, such as technology and preferences, or in terms of scale, composition, and technique effects, also known as “proximate factors.” Scale refers to the effect of an increase in the size of the economy, holding the other effects constant, and should increase environmental impacts. The composition and technique effects must outweigh this scale effect for pollution or other environmental impacts to fall in a growing economy. The composition effect refers to the economy’s mix of different industries and products, which differ in pollution intensities. Finally, the technique effect refers to the remaining change in pollution intensity. This will include contributions from changes in the input mix, for example, substituting natural gas for coal; changes in productivity that result in less use, ceteris paribus, of polluting inputs per unit of output; and pollution control technologies that result in less pollutant being emitted per unit of polluting input. Over the course of economic development, the mix of energy sources and economic outputs tends to evolve in predictable ways. Economies start out mostly agricultural, and the share of industry in economic activity first rises and then falls as the share of agriculture declines and the share of services increases. We might expect the impacts associated with agriculture, such as deforestation, to decline, and naively expect the impacts associated with industry, such as pollution, would first rise and then fall. However, the absolute size of industry rarely does decline, and it is improvement in productivity in industry, a shift to cleaner energy sources, such as natural gas and hydro-electric power, and pollution control that eventually reduce some industrial emissions. Offshoring of pollution probably plays only a small role in cutting emissions in developed economies (Kander et al., 2015). Static theoretical economic models of deep determinants that do not try to also model the economic growth process can be summarized in terms of two parameters: the elasticity of substitution between dirty and clean inputs, which summarizes how difficult it is to cut pollution; and the elasticity of the marginal utility of consumption with respect to consumption, which summarizes how hard it is

The environmental Kuznets curve  239

to increase consumer well-being with more consumption (Pasten and Figeroa, 2012). It is usually assumed that these consumer preferences are translated into policy action. Pollution is then more likely to increase as the economy expands, the harder it is to substitute other inputs for polluting ones and the easier it is to increase consumer well-being with more consumption. If these parameters are constant, then pollution either rises or falls with economic growth. Only if they change over time will pollution first rise and then fall. The various theoretical models can be classified as ones where the EKC is driven by changes in the elasticity of substitution as the economy grows or models where the EKC is primarily driven by changes in the elasticity of marginal utility. Dynamic models that model the economic growth process alongside changes in pollution are harder to classify. The Green Solow model developed by Brock and Taylor (2010) explains changes in pollution as a result of the competing effects of economic growth and a constant rate of improvement in pollution control. Fast-growing middle-income countries, such as China, then have rising pollution, and slower-growing developed economies have falling pollution. An alternative model developed by Ordás Criado et al. (2011) also suggests that pollution rises faster in faster-growing economies but that there is also convergence between countries with initially high or low levels of pollution so that countries with higher levels of pollution tend to reduce pollution faster than do countries with low levels of pollution.

Beyond the environmental Kuznets curve

Some recent empirical research (e.g. Ordás Criado et al., 2011; Stern et al., 2017) builds on these dynamic models to provide a richer interpretation of the trends in environmental impacts and economic activity than conventional EKC studies do (Stern, 2017). We can distinguish between the effects of economic growth and the level of GDP per capita on environmental impacts. We can also distinguish between the effects of economic growth and the simple passage of time. Economic growth usually increases environmental impacts, but the size of this effect varies across impacts and the impact

of growth often declines as countries get richer. Time effects reduce many impacts at all levels of GDP per capita. However, richer countries often make more rapid progress in reducing environmental impacts. This is the effect of the level of GDP per capita. Rapid economic growth in middle-income countries, such as China or India, is more likely to overwhelm the time effect in those countries, as suggested by Brock and Taylor (2010). Finally, there is often convergence among countries, so that those that have relatively high levels of impacts reduce them more rapidly or increase them more slowly than countries with low levels of impacts. These combined effects explain more of the variation in pollution emissions or concentrations than either the classic EKC model or models that assume that either only convergence or growth effects alone are important. Therefore, while being rich means a country might do more to clean up its environment, getting rich is likely to be environmentally damaging. David I. Stern

References

Arrow, K., Bolin, B., Costanza, R., Dasgupta, P., Folke, C., Holling, C. S., Jansson, B.-O., Levin, S., Mailer, K.-G., Perrings, C., Pimentel, D., 1995. Economic growth, carrying capacity, and the environment. Science 268, 520–21. Beckerman, W., 1992. Economic growth and the environment: Whose growth? Whose environment? World Development 20, 481–96. Brock, W. A., Taylor, M. S., 2010. The Green Solow model. Journal of Economic Growth 15, 127–53. Daly, H. E., 1993. The perils of free trade. Scientific American 269(November), 50–57. Grossman, G. M., Krueger, A. B., 1991. Environmental impacts of a North American Free Trade Agreement. NBER Working Papers 3914. Kander, A., Jiborn, M., Moran, D. D., Wiedmann T. O., 2015. National greenhouse-gas accounting for effective climate policy on international trade. Nature Climate Change 5, 431–5. Ordás Criado, C., Valente, S., Stengos, T., 2011. Growth and pollution convergence: Theory and evidence. Journal of Environmental Economics and Management 62, 199–214. Panayotou, T., 1993. Empirical tests and policy analysis of environmental degradation at different stages of economic development. Working Paper 238, Technology and Employment

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240  Elgar encyclopedia of ecological economics Programme, International Labour Office, Geneva. Pasten, R., Figueroa, E., 2012. The environmental Kuznets curve: A survey of the theoretical literature. International Review of Environmental and Resource Economics 6, 195–224. Stern, D. I., 2005. Beyond the environmental Kuznets curve: Diffusion of sulfur-emissions-abating technology. Journal of Environment and Development 14(1), 101–24. Stern, D. I., 2017. The environmental Kuznets curve after 25 years. Journal of Bioeconomics 19, 7–28. Stern, D. I., Common, M. S., Barbier, E. B., 1996. Economic growth and environmental degra-

David I. Stern

dation: the environmental Kuznets curve and sustainable development. World Development 24, 1151–60. Stern, D. I., Gerlagh, R., Burke, P. J., 2017. Modeling the emissions–income relationship using long-run growth rates. Environment and Development Economics 22(6), 699–724. World Bank, 1992. World Development Report 1992: Development and the Environment. Oxford: Oxford University Press. World Commission on Environment and Development, 1987. Our Common Future. Oxford: Oxford University Press.

40. Environmental limits Concepts and definitions

Limits have long history. Their first conceptualization is attributed to Archimedes of Syracuse in the third century BC, who developed the notion to measure curved figures and the volume of a sphere. Thereafter the concept has been widely used in mathematics, later making its way into philosophical, legal, economic, and environmental debates. The Oxford English Dictionary defines limits as ‘a point or level beyond which something does not or may not extend’, or as ‘a restriction on the size or amount of something permissible or possible’. The Cambridge Dictionary similarly defines limits as ‘the greatest amount, number, or level of something that is either possible or allowed’ (emphasis added). Hence, the idea of limits can be used in both a descriptive and a normative sense, describing physical properties as much as norms and regulations defining where limits ought to be set. Environmental limits describe critical points at which pressure on a natural resource or ecosystem creates abrupt or irreversible change (Haines-Young et al. 2006). In scholarly and policy discourse environmental limits has been used as a generic term for a broad range of related concepts and ideas, such as ecological limits, biophysical limits, resource limits, ecological thresholds, tipping points, and planetary boundaries, concepts which are sometimes used loosely and interchangeably in the literature. Since the late 1960s environmental limits were often discussed in relation to ‘carrying capacity’ (the maximum population an ecosystem can sustain), in particular among Malthusian environmentalist strands concerned with population limits (e.g. Ehrlich 1968). This approach lost traction after the 1980s, faced with criticism for neglecting other pressures to environmental limits such as technology and consumption (Commoner 1971). In recent decades, discussions on limits have drifted towards notions of ecological thresholds, tipping points, and regime shifts, all of which describe abrupt ecological transitions ranging from population collapse to shifts at the whole ecosystem scale (Dakos 2019). A key insight of these concepts is that ecosystems do not respond

to pressures in gradual or predictable ways; when thresholds are crossed, ecosystems can react in non-linear ways, sometimes resulting in sudden reorganizations or even collapse (Holling 1973, Scheffer et al. 2001, Lenton 2013). At global scale, limits are increasingly called ‘planetary boundaries’, an idea that defines the ecological confines within which humanity can safely operate. Scientists claim that several planetary boundaries, such as biosphere integrity and biogeochemical flows, have already been transgressed (Steffen et al. 2015) An important distinction, often blurred in the environmental and economic literatures, concerns limits and scarcity (Mehta 2013; Gómez-Baggethun 2022a). Scarcity defines what we cannot use as much as we want and is a relation between means and ends; physical limits relate to what exists in finite quantities and associated regulations. The distinction matters because something that is limited is not necessarily scarce (e.g. O2 in the atmosphere), but also because a failure to make this distinction can mask processes where scarcity is artificially created (e.g. through privatization and enclosure of land; Turner 1993).

Classifying environmental limits

Just as limits in the broader sense, environmental limits can be descriptive or normative. Hence, a distinction is to be made between ‘natural limits’ (e.g. the remaining amount of oil on Earth) and ‘regulatory limits’ (e.g. the maximum amount of oil that can be legally extracted under a given governance regime; Jax 2016; Gómez-Baggethun 2020). While often neglected (e.g., Robbins, 2020), the key distinction between ‘natural’ and ‘regulatory’ limits have long been made in environmental policy and governance. For example, ecological economics precursor Ciriacy-Wantrup (1968) distinguished ‘critical zones’ from ‘safe minimum standards’, just as recent work by Earth system scientists distinguish ‘ecological thresholds’ from the planetary boundaries defining ‘safe operating space’ for humanity (Rockström et al. 2009). Ecological thresholds are defined by physical parameters, whereas regulatory limits are negotiated and contingent on, for example, risk aversion, existing knowledge, and environmental awareness (Gómez-Baggethun

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2022b). As opposed to natural limits describing physical properties of the environment, normative limits are by definition subjective and constructed. Descriptive accounts of limits, of current use in ecology and the environmental sciences, focus on observable biophysical properties or phenomena. Some such limits, like the Earth’s surface (510 million km2) and its volume (abut 1 trillion km3) are trivial, whereas others, such as climate tipping points at which catastrophic effects shall be expected (e.g. the 1.5 degrees defined by the Paris climate agreement) are subject to large uncertainties (Dudney and Suding 2020). Used in a descriptive sense, environmental limits can in turn be divided in two sub-categories: ‘resource limits’, which define finite stocks of resources, and ‘system limits’, such as tipping points in ecosystems (Green 2021). Normative limits define the level at which the limit is posited, either through social norms or formally sanctioned rules (Vatn 2016). They cover a diverse set of ideas raging from ethics of self-restraint to legally binding regulations. Regulatory limits, also known as decision thresholds or management thresholds, refer to points in some variable condition to which a risk of system change is permitted or accepted (Johnson 2013). These limits act as institutions for protecting ecosystem and resource systems by setting limits on permitted amounts of pollution or resource extraction (Vatn 2015). For example, in the fisheries literature, ‘precautionary limits’ or ‘reference points’ are set to ensure that irreversible harm does not occur to economically or ecologically important species (Haines-Young et al. 2006). While a typology of limits is useful for analytical purposes, no sharp boundaries always apply between the above-described categories. Overlaps exist, and the common usage of the term environmental limits often merges normative and descriptive considerations. One of the most influential approaches combining normative and descriptive accounts of limits is the doughnut economics framework, which defines a just and safe operating space for humanity by means of integrating an ‘ecological ceiling’ (planetary boundaries) and a social foundation (defined from internationally agreed minimum social standards such as the Sustainable Development Goals; Raworth 2017). Erik Gómez-Baggethun

The case for limits to growth

The idea that the economy cannot grow beyond the physical limits of the biosphere is the central proposition of ecological economics (Daly 2014). Ideas of limits to growth are often traced back to Malthus’s (1798) concerns on population growth outpacing food production, but notions of resource limits are present in the work of the physiocrats and all major classical economists, including Ricardo’s thesis on the decreasing returns on land, Marx’s notion of the ‘metabolic rift’, and Mill’s precursory writings on a stationary state economy (Gómez-Baggethun and Naredo 2015). In environmental and policy debates the thesis of limits gained momentum 50 years ago with the publication of the Club of Rome report ‘Limits to Growth’ (Meadows et al. 1972). The case for physical limits is often grounded in the thermodynamic vision of the economy, first theorized by ecological economics father Georgescu-Roegen (1971) and later elaborated by Daly (2014). This vision, sometimes referred to as the ‘nested economy’ framework, portrays the economy as a subsystem of the biosphere, where the economy depends on ecosystems as both a source of resources and a sink of waste. Industrial metabolism transforms energy and materials into goods and services in a process that irreversibly converts (low entropy) stocks of resources into (high entropy) waste. Physical resources are finite, and entropy prevents complete recycling. Consequently, the thesis goes, the economy cannot grow forever. In our finite planet, sustainable limits are defined by the regenerative and assimilative capacities of ecosystems (Daly 2014). Beyond a certain scale, the economy enters in conflict with such limits, the social and environmental costs of growth accelerate (Mishan 1967, Kapp 1978), and environmental conflicts multiply at the extraction frontiers and in relation to transport and waste disposal (Martínez-Alier 2014). The thesis of limits achieved its splendour in the 1970s but thereafter declined in response to criticism from ecomodernists and critical social scientists (Gómez-Baggethun 2022b). Fifty years since the publication of ‘Limits to Growth’, however, limits have made a comeback to political debates with renovated force (Dobson 2016, Benjaminsen 2021).

Environmental limits  243

Contested claims and controversies

Limits are a highly contentious concept. Attacked and defended with equal passion, they define major ontological and ideological divides, and have been scenario of enduring confrontations across disciplines (Gomez-Baggethum, 2022). Natural scientists warn about catastrophic effects of crossing tipping points, neoclassical economists claim that technological innovation and resource substitution make physical limits of little relevance in the foreseeable future (a vision shared by many politicians and business representatives), and critical social scientists claim that limits serve the political agendas of conservation elites. Controversies of limits cover a broad range of topics, spanning from their ontology to their practical relevance, and political implications. Regarding the ontology of limits, tensions have become apparent between those who understand limits as objective properties of the environment and those who emphasize their subjective and constructed nature. The first vision is common in natural sciences, such as ecology and Earth system science, where scientists claim that limits can be described and identified (under given uncertainties) through physical parameters (e.g. Rockström et al. 2009). The latter has stronger currency among social and political scientists, who emphasize their relative and socially constructed character (e.g. Kallis 2020). Second, regarding the practical implication of limits, strong divisions arise between those who claim that limits are already being transgressed, involving risks for sudden ecosystem change or even collapse (Rockström et al. 2009, Steffen et al. 2015), and those who claim that technology and resource substitution render limits irrelevant in the foreseeable future (Barnett and Morse 1963, Solow 1973, Stiglitz 1979), including those that dismiss limits as unfounded catastrophism (Beckerman 1974, Bruckner 2013). Emphasis on limits to growth is strong among ecological economists (Daly 2014, Jackson 2017, Gómez-Baggethun 2020), whereas neoclassical economists and ecomodernists tend to deny that limits have any immediate relevance (Solow 1973, Stiglitz 1979, Simon 1981, Asafu-Adjaye et al. 2015). Finally, tensions arise with regard to the political implications of limits, and whether

discourses on limits serve reactionary vs emancipatory agendas. Political ecologists and critical social scientists have attacked (sometime correctly) the case for limits as Malthusian discourses that diverge the blame for environmental destruction from the rich and powerful to the poor and marginalized (Benjaminsen et al. 2015), used by ‘expertocracies’ to impose top-down population policies in the name of objective science (Gorz 1993, Harvey 1974). However, ecological economists and degrowth proponents have argued that, whereas it is true that the case of limits does not necessarily involve an endorsement of Malthusian ideology, noting that ideas of limits are also mobilized for emancipatory political agendas with emphasis on equity and justice (Kallis 2020, Gómez-Baggethun 2020, 2022a). While reckoning population to be an important variable, various environmentalist strands have long argued against Malthusians like Hardin (1968) and Ehrlich (1968), setting the focus on technology and affluence (Commoner 1971), highlighting differentiated responsibilities for ecological damage, and noting that growth among the rich occurs at the expense environmental destruction and the poor (Martínez-Alier 2014).

Conclusions

Limits to growth is a central proposition of ecological economics. As an interdisciplinary field of knowledge, ecological economics pays attention to both descriptive and normative accounts of environmental limits, taking due consideration of both their biophysical and socio-political dimensions. Besides due consideration of empirical research on ecological thresholds, tipping points, and planetary boundaries as much as philosophical and political examinations of how physical limits intersect with considerations of justice, equity, and power. This requires differentiating socioenvironmental problems characterized by objective physical scarcities from situations where scarcity is artificially created through mechanisms of exclusion and unequal resource appropriation through commodification and enclosure.

Acknowledgements

This research received partial funding from NMBU’s Sustainability Arena ‘Embedding Erik Gómez-Baggethun

244  Elgar encyclopedia of ecological economics

Planetary Boundaries in Science Policy and Education’ (Grant Nr. 1850092016AA). Erik Gómez-Baggethun

MA: Edward Elgar Publishing. Gómez-Baggethun, E. (2022b). Political ecological correctness and the problems of limits. Political Geography, 98, 102622. Gómez-Baggethun, E., and Naredo, J. M. (2015). In search of lost time: The rise and fall of limits References to growth in international sustainability policy. Asafu-Adjaye, J., Blomquist, L., Brand, S., Sustainability Science, 10(3), 385–95. Brook, B. W., De Fries, R., Ellis, E., et al. Gorz, A. (1993). Political ecology: Expertocracy (2015). An ecomodernist manifesto. https://​ versus self-limitation. New Left Review, 202, thebreakthrough​.org/​manifesto/​manifesto​ 55–67. -english Green, F. (2021). Ecological limits: Science, Barnett, H. J., and Morse, C. (1963). Scarcity and justice, policy, and the good life. Philosophy growth: The economics of natural resource Compass, 16(6), e12740. availability. Natural Resources Journal, 3(3), Haines-Young, R., Potschin, M., and Cheshire, D. 550. (2006). Defining and identifying environmental Beckerman, W. (1974). In defence of economic limits for sustainable development. A scoping growth (Vol. 1976). London: J. Cape. study. Final full technical report to Defra. Benjaminsen, T. A. (2021). Virtual forum introCentre for Environmental Management School duction: Environmental limits, scarcity and of Geography, University of Nottingham, degrowth. Political Geography, 87, 102344. Nottingham. https://​www​.nottingham​.ac​.uk/​ Benjaminsen, T. A., Reinert, H., Sjaastad, E., cem/​pdf/​NR0102​_FTR​_Final​.pdf and Sara, M. N. (2015). Misreading the Arctic Hardin, G. (1968). The tragedy of the commons: landscape: A political ecology of reindeer, carThe population problem has no technical solurying capacities, and overstocking in Finnmark, tion; it requires a fundamental extension in Norway. Norwegian Journal of Geography, morality. Science, 162(3859), 1243–8. 69(4), 219–29. Harvey, D. (1974). Population, resources, and Bruckner, P. (2013). The fanaticism of the apocthe ideology of science. Economic geography, alypse: Save the Earth, punish human beings. 50(3), 256–277. Malden, MA: Polity Press. Holling, C. S. (1973). Resilience and stability of Ciriacy-Wantrup, S. V. (1968). Resource conecological systems. Annual Review of Ecology servation: Economics and policies. Berkeley: and Systematics, 4(1), 1–23. University of California Press. Jackson, T. (2017). Prosperity without growth. Commoner, B. (1971). The closing circle: Nature, London: Earthscan. man, and technology. New York: Alfred A. Jax, K. (2016). Thresholds, tipping points and Knopf. limits. In M. Potschin and K. Jax (Eds.), Dakos, V. (2019). Ecological transitions: Regime OpenNESS ecosystem services reference book. shifts, thresholds and tipping points. Oxford: EC FP7 Grant Agreement no. 308428. https://​ Oxford University Press. www​.guidetoes​.eu/​synthesispapers/​OpenNESS​ Daly, H. E. (2014). Beyond growth: The econom_SP4​_Thresholds​_2016​.pdf ics of sustainable development. Boston: Beacon Johnson, C. J. (2013). Identifying ecological Press. thresholds for regulating human activity: Dobson, A. (2016). Are there limits to limits? Effective conservation or wishful thinking? In T. Gabrielson, C. Hall, J. M. Meyer, and Biological Conservation, 168, 57–65. D. Schlosberg (Eds.), The Oxford handbook Kallis, G. (2020). Limits: Why Malthus was of environmental political theory, 289–303. wrong and why environmentalists should care. Oxford: Oxford University Press. Stanford, CA: Stanford University Press. Dudney, J., and Suding, K. N. (2020). The elusive Kapp, W. (1978). The social costs of business search for tipping points. Nature Ecology & enterprise. Nottingham, UK: Spokesman Evolution, 4, 1449–50. Books. Ehrlich, P. R. (1968). The population bomb. New Lenton, T. M. (2013). Environmental tipping York: Ballantine Books. points. Annual Review of Environment and Georgescu-Roegen, N. (1971). The entropy law Resources, 38, 1–29. and the economic process. Cambridge, MA: Malthus, T. (1798). An essay on the principle of Harvard University Press. population as it affects the future improvement Gómez-Baggethun, E. (2020). More is more: of society, with remarks on the speculations of Scaling political ecology within limits to Mr. Goodwin, M. Condorcet and Other Writers growth. Political Geography, 76, 102095. (1st ed.). London: J. Johnson in St Paul’s Gómez-Baggethun, E. (2022a). Limits. In L. Church-yard. Pellizzoni, M. Leonardi, and V. Asara (Eds.), Martínez-Alier, J. (2014). The environmentalism Handbook of critical environmental politics, of the poor. Geoforum, 54, 239–41. 129–40. Cheltenham, UK, and Northampton, Meadows, D. H., Meadows, D. L., Randers, J., and

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Environmental limits  245 Behrens, W. W. (1972). The limits to growth. A report for the Club of Rome’s Project on the Predicament of Mankind. New York: Universe Books. Mehta, L. (2013). The limits to scarcity: Contesting the politics of allocation. New York: Routledge. Mishan, E. J. (1967). The costs of economic growth (Vol. 9). London: Staples Press. Raworth, K. (2017). Doughnut economics: Seven ways to think like a 21st-century economist. White River Junction, VT: Chelsea Green Publishing. Robbins, P. (2020). Is less more . . . or is more less? Scaling the political ecologies of the future. Political Geography, 76, 102018. Rockström, J., Steffen, W., Noone, K., Persson, A. A., Chapin, F. S. III, et al. (2009). A safe operating space for humanity. Nature, 461, 472–5. Scheffer, M., S. Carpenter, J. A. Foley, C. Folke, and B. Walker. (2001). Catastrophic shifts in ecosystems. Nature, 413(6856), 591–6. Simon, J. L. (1981). The ultimate resource. Princeton, NJ: Princeton University Press.

Solow, R. M. (1973). Is the end of the world at hand? Challenge, 2, 39–50. Steffen, W., Richardson, K., Rockstrom, J., Cornell, S. E., Fetzer, I., and Bennett, E. M. (2015). Planetary boundaries: Guiding human development on a changing planet. Science, 347(6223), 1259855. Stiglitz, J. E. (1979). A neoclassical analysis of the economics of natural resources. In V. K. Smith (Ed.), Scarcity and growth reconsidered, 36–66. Washington, D.C.: Resources for the Future. Turner, M. (1993). Overstocking the range: A critical analysis of the environmental science of Sahelian pastoralism. Economic Geography, 69(4), 402–21. Vatn, A. (2015). Environmental governance: Institutions, policies and actions. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Vatn, A. (2016). On limits. In J. Farley and D. Malghan (Eds.), Beyond uneconomic growth, 83–105. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing.

Erik Gómez-Baggethun

41. Environmental stewardship Environmental stewardship is a popular way of conceiving of morally decent and responsible human conduct toward the natural world. The word “steward” is derived from the Old English “stigwaerd,” which referred to servants in charge of farms, manors, or landed estates. Over the centuries, its meaning has expanded to include people engaged in a much wider range of occupations and activities (e.g., wine steward, ship steward, data steward). What makes these all instances of stewardship is that, in each case, individuals act as trustees caring for things and persons in the best interests of someone other than themselves. The concept of environmental stewardship takes this expansion of meaning a step further. Like other forms of stewardship, environmental stewardship is a form of trusteeship of goods, resources, or persons, in the interests of others. But the scope of environmental stewardship is expanded to include the whole of the natural world. Eligibility to enact the role has likewise expanded to include individuals and groups acting voluntarily, nonprofit organizations, and governmental and nongovernmental organizations, as well as the individuals specifically employed to perform stewardship activities. Thus the concept of environmental stewardship overlaps with but is distinct from related concepts of environmental management and sustainable development. Environmental management, or “wise use” of environmental resources, is an activity constrained by prudent regard for the limits of environmental resources or ecosystem resources in the long-term best interests of the managing individual or group. Prudent regard for the long term has been a motivating factor behind some individuals’ and organizations’ participation in voluntary environmental management programs created by governments and nonprofit organizations to reduce their environmental impact beyond the requirements of local, national, or international regulation. Sometimes called “stewardship” programs, voluntary environmental programs reward reductions of pollutants, overharvesting, or energy inefficiencies with certifications that make the participants’ products and

services more attractive to consumers and/ or prevent new or increased governmental regulations (Tashman et al., 2022). Examples include the Forest Stewardship Council, the Marine Stewardship Council, Leadership in Energy and Environment Design (LEED), the International Organization for Standardization (ISO), and Responsible Care® (American Chemistry Council). Such programs further goals that environmental stewards share, but because these programs rely on individual or corporate self-interest to motivate participants, their participants do not act as environmental stewards. Sustainable development, by contrast, incorporates an altruistic (other-directed) concern for the good of our descendants. The goals of sustainable development are to manage the practices on which human beings rely for food, shelter, education, recreation, health care, and cultural goods to ensure those practices can be equitably maintained for generations into the future. As many of these practices draw upon or impact important ecosystem services, sustainable development projects often seek to reduce or restrain these impacts so that they do not become destabilizing. However, as a review of the United Nation’s 17 sustainable development goals will reveal, sustainable development activities and projects are more wide-ranging than those of environmental stewardship. These initiatives aim to ensure human practices are sustainable economically, socially, politically, and culturally, as well as ecologically. The conservation of the natural environment and ecosystem services is only one of several forms that sustainable development activities can take. By contrast, environmental stewardship is solely focused on human environmental impacts and adopted from a concern for future generations of non-human as well as human life. Religious and spiritual traditions that represent the world as entrusted to the care of human beings by the world’s Creator have played a role in motivating stewardship activities by their followers. Others are motivated by love for nature, appreciation of natural beauty, and/or by ethical outlooks according to which non-human life and ecosystems as wholes possess intrinsic moral value that merits respect. Biocentrism, the view that life is inherently morally valuable, is one such outlook. Aldo Leopold’s (1949) ecocentric “land ethic,” according to which humans are

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Environmental stewardship  247

enjoined to view themselves as citizens of biotic as well as human communities, owing obligations to both, has been another widely influential Western source. Indigenous kincentric ecological outlooks according to which human beings aren’t merely fellow citizens but (more or less distant) kin to the non-human members of their ecological communities are perhaps even more significant contributors worldwide (Romero-Briones et al., 2020; Wehi et al., 2021). Some environmentalists have criticized the conception as flawed on the ground that the original concept arose in hierarchical and elitist societies. Others presume stewards will necessarily sacrifice the interests of the natural world whenever they conflict with those of human beings. Neither is well founded. We do not reject democracy as a political concept because it arose in an elitist and xenophobic society. And since there are environmental stewards motivated by biocentrism, ecocentrism, and kincentric ecological beliefs, environmental stewards are not necessarily committed to narrowly anthropocentric goals. Recently environmental stewardship has been the subject of sociological, psychological, and economic research with the goal of determining whether promoting stewardship as a moral ideal could increase public support for and engagement in environmental initiatives. To this end, researchers are investigating the motivations of different groups, the

resources needed to mount effective projects, and methods for measuring and comparing their outcomes (Bennett et al., 2018). Jennifer Welchman

References

Bennett, N.J., Whitty, T.S., Finkbeiner, E., Pittman, J., Bassett, H., Gelcich, S., and Allison, E.H. 2018. Environmental Stewardship: A Conceptual Review and Analytical Framework. Environmental Management 61 (4): 597–614. Leopold, Aldo. 1949. A Sand County Almanac. Oxford University Press. Romero-Briones, A-dae, Salmon, Enrique, Renick, Hillary, and Costa, Temra. 2020. Recognition and Support of Indigenous California Land Stewards, Practitioners of Kincentric Ecology. First Nations Development Institute and California Foodshed Funders. https://​cerestrust​ .org/​wp​-content/​uploads/​Indigenous​-California​ -Land​-Stewards​-Practitioners​-of​-Kincentric​ -Ecology​-Report​-2020​.pdf Tashman, Peter, Flankova, Svetlana, van Essen, Marc, and Marano, Valentina. 2022. Why Do Firms Participate in Voluntary Environmental Programs? A Meta-Analysis of the Role of Institutions, Resources, and Program Stringency. Organization & Environment 35(1): 3–29. https://​doi​.org/​10​.1177/​1086026621990063 Wehi, P.M., van Uitregt, V., Scott, N.J., et al. 2021. Transforming Antarctic Management and Policy with an Indigenous Māori Lens. Nature Ecology & Evolution 5: 1055–9. https://​doi​.org/​ 10​.1038/​s41559​-021​-01466​-4

Jennifer Welchman

42. Environmental tax reform

However, the revenues could be used in a number of other ways, including:

Environmental taxation (qv) entails the imposition of a tax on a substance, or plausible proxy of a substance, that has a specific negative environmental impact. Environmental tax reform (ETR) involves environmental taxation, but with specific consideration of the use of the revenues from the environmental tax in the context of potential changes to the tax system as a whole. There is no consensus on what use of the revenues is required for an imposition of environmental taxes to qualify as an ETR. A common understanding is that ETR entails a tax shift, rather than an overall increase in the tax burden, and ETR has been defined explicitly in these terms as ‘a reform of the national tax system where there is a shift of the burden of taxation from conventional taxes, for example on labour, to environmentally damaging activities, such as resource use or pollution’ (European Environment Agency [EEA], 2006, p. 84) – that is, a shift from taxing ‘goods’ to taxing ‘bads’. Where ETR involves the reduction in labour and social security taxes, there has been much speculation that it could result in a double dividend (qv), whereby environmental impacts are reduced and gross domestic product (GDP) increased through the reduction in labour market distortions arising from labour taxation.

Source:

● Investments in innovation relating to clean technologies (also called eco-innovation; e.g. EEA, 2012a). ● Compensation for low-income households, to remove any regressive effects from the ETR (e.g. EEA, 2012b). ● Lump-sum returns to households (e.g. Government of British Columbia [GBC], 2021a, 2021b, which operates a sliding scale of lump-sum payments depending on household income). ● Maintenance of the competitiveness of affected industries (this mainly takes the form of giving tax rebates to energy-intensive sectors, sometimes requiring commitments to increase energy efficiency, for example, the UK Climate Change Agreements; see Her Majesty’s Government, 2021). EEA (2006, p. 84) distinguishes between ETR and environmental (or ecological) fiscal reform (EFR), which it defines as ‘a broader approach, which focuses not just on shifting taxes and tax burdens, but also on reforming economically motivated subsidies, some of which are harmful to the environment)’. However, the Organisation for Economic Co-operation and Development (OECD, 2017, p. 6) identifies three approaches that fall under the name of EFR. The broadest is ‘a range of taxation or pricing instruments that can raise revenue, while simultaneously furthering environmental goals’, which says nothing about use of the revenues. Narrower is ‘a tax shift

Andersen and Ekins (2009, Figures 7.6 and 7.7).

Figure 42.1 The effect of ETR on seven EU countries in respect of greenhouse gas emissions (left panel) and GDP (right panel)

248

Environmental tax reform  249 Table 42.1 ETR scenarios and their impacts Scenario

LS1

HS1

HS2

Energy price

BL

BH

BH

HS3 BH

CO2/GHG reduction in

−15%/−20%

−15%/−20%

−15%/−20%

−25%/−30%

Materials tax

15%

15%

15%

15%

Revenue recycling (in

Employment and income Employment and income Low-carbon investment, Employment and income

different proportions in

taxes

taxes

Other

  

  

  

International cooperation

Impacts

  

  

Productivitya

 

 

    

    

(GINFORS)

 

 

 

 

 Material

1.97

0.91

0.84

1.78

 Labour

−3.02

−0.93

−0.71

−2.61

 Carbon Carbon prices (2008

17.17  

8.59  

8.99  

21.35  

euros)

 

 

 

 

 E3ME

142

59

53

204

 GINFORS GDPa

120

68

61

184

 

 

 

 

 E3ME

0.6

0.2

0.8

0.5

 GINFORS Employmenta

−3.0

−0.6

−0.3

−1.9

 

 

 

 

 E3ME

2.2

1.1

1.1

2.7

 GINFORS Inflationa

0.0

0.4

0.4

0.8

 

 

 

 

  E3ME (price level)

1.6

0.8

0.7

1.8

  GINFORS (CPI)

3.0

0.9

1.1

4.1

2020 (from 1990 level)

employment, and income taxes taxes

different scenarios)

Note: a Results are percentage difference from BL in 2020. Source: Ekins and Speck (2011, with figures extracted from Chapter 9).

from labour towards environmental use, supplemented by the reform or removal of environmentally adverse subsidies’, which is close to the EEA’s definition of EFR cited above. In between is EFR, which ‘is frequently discussed as a means of bringing about a so-called “tax shift” in which a progressive increase in the revenues generated through environmentally related taxes provides a rationale for reducing taxes derived from other sources, such as income, profits and employment, the taxation of which is less desirable’ (OECD, 2017, p. 6), which closely resembles the EEA definition of ETR. OECD (2017, p.7) itself opts for EFR involving ‘(a) environmental policy using market-based instruments to reflect the cost of environmental damage in prices faced by

polluters and (b) raising public revenue and deploying it in a socially useful way’. ETR has been most commonly implemented in Europe, and the main tax base involved has been carbon and/or energy. A detailed ex post evaluation of the seven main ETRs carried out in Europe between 1990 and 2005 (Andersen and Ekins, 2009) concluded that all but one (Slovenia, where the tax reform was very small) had been successful in reducing greenhouse gas (GHG) emissions by up to 6 per cent; that they had resulted in a small increase in GDP for the countries concerned, over what it would have been without the ETR (Figure 42.1); and that six of the seven ETR countries had performed slightly better than the nine countries in the then European Union (EU-15) that had not implemented ETRs. Paul Ekins

250  Elgar encyclopedia of ecological economics

An ex ante assessment of a major ETR in Europe was reported in Ekins and Speck (2011). Four scenarios were modelled using two kinds of global models, a macro-econometric model (E3ME) and a computable general equilibrium model (GINFORS). The six scenarios had two baselines, one with a low (BL) and one with a high (BH) energy price, and four ETR scenarios involving both a carbon and a materials tax, and full revenue recycling. The details of the scenarios are set out in Table 42.1, together with their results in respect of productivity, carbon prices, GDP, employment, and inflation. The major points of the study’s results may be highlighted as follows: ● GHG emissions: the carbon tax was levied at the rates in each scenario necessary to reach the then targets in the EU. The tax rates need to be higher with low oil prices (LS1) or greater emission reduction (HS3). When the HS3 carbon tax is levied globally, global CO2 emissions are stabilised between 2010–20 at about 29 billion tonnes of CO2, whereas in the baseline they reach more than 34 billion tonnes (Ekins & Speck, 2011, Figure 11.7, p. 304; the actual global CO2 level in 2019, before the pandemic, was 33.4 billion tonnes1). ● Carbon prices in HS1 and HS2 in 2020 in both models are close to current EU ETS levels. They are higher when oil prices are low (LS1) or the required GHG emission reduction is higher (HS3). ● The macro-economic effects in the scenarios are small: ● GDP rises above the baseline in one model, and falls in another, with the largest effect being a 3 per cent reduction in GDP in the low oil price scenario, LS1 (GINFORS). However, in all scenarios the effect on GDP growth is small, and throughout the period the European economy continues to grow at around 2.0 per cent or more, compared to the baseline level of 2.2 per cent. ● Employment rises slightly in both models, by more than GDP in E3ME, so that labour productivity falls in both models. Paul Ekins

● The tax shift is slightly inflationary (a maximum of around 0.3 per cent per year over the period, in LS1 and 0.4 per cent per year in HS3 for the larger GHG reduction [GINFORS]). Clearly the results of a single study are illustrative rather than definitive, but they suggest that ETR could play a useful role in helping European and other countries meet the much more stringent GHG reduction targets implied by the Paris Agreement temperature target, which are now deemed necessary to avoid the worst effects of climate change. Paul Ekins

Note 1.

See https://​www​.iea​.org/​articles/​global​-energy​-rev iew​-co2​-emissions​-in​-2020.

References

Andersen, M.S., and Ekins, P. (Eds.) 2009 Carbon-Energy Taxation: Lessons from Europe, Oxford University Press, Oxford. Ekins, P., and Speck, S. (Eds.) 2011 Environmental Tax Reform: A Policy for Green Growth, Oxford University Press, Oxford. European Environment Agency (EEA) 2006 Market-based instruments for environmental policy in Europe, EEA Technical Report No 8/2005, EEA, Copenhagen, https://​www​ .eea​.europa​.eu/​publications/​technical​_report​ _2005​_8 EEA 2012a Environmental tax reform in Europe: opportunities for eco-innovation, EEA Technical Report No 17/2011, EEA, Copenhagen, https://​www​.eea​.europa​.eu/​ publications/​environmental​-tax​-reform​ -opportunities EEA 2012b Environmental tax reform in Europe: implications for income distribution, EEA Technical Report No 16/2011, EEA, Copenhagen, https://​www​.eea​.europa​.eu/​ publications/​environmental​-tax​-reform​-in​ -europe Government of British Columbia (GBC) 2021a British Columbia’s carbon tax, https://​www2​ .gov​.bc​.ca/​gov/​content/​environment/​climate​ -change/​clean​-economy/​carbon​-tax GBC 2021b Climate action tax credit, https://​ www2​.gov​.bc​.ca/​gov/​content/​taxes/​income​ -taxes/​personal/​credits/​climate​-action Her Majesty’s Government 2021 Climate Change Agreements, https://​www​.gov​.uk/​guidance/​ climate​-change​-agreements​-​-2 Organisation for Economic Co-operation

Environmental tax reform  251 and Development (OECD) (2017, June) Environmental fiscal reform: progress, prospects and pitfalls, Report for the G7 Environment Ministers, OECD, Paris, https://​

www​.oecd​.org/​tax/​tax​-policy/​environmental​ -fiscal​-reform​-G7​-environment​-ministerial​ -meeting​-june​-2017​.pdf

Paul Ekins

43. Environmental taxation and the double dividend Introduction

As a policy instrument for the control of pollution, environmental taxation or emissions charges will reduce pollution because firms or individuals will reduce emissions in order to avoid paying the tax. Under a range of market conditions, a pollution tax will generally be more cost-effective at reducing pollution than regulations: the total abatement cost of achieving a specified level of pollution reduction will generally be lower under a pollution tax than for a regulatory “command-and-control” approach. This social cost advantage, well-established in both theory and empirical studies, is largely due to the flexible way in which environmental taxes allow for decentralized decisions at the firm or household level about how, and how much, to reduce pollution given their costs and opportunities for doing so. The idea that taxation can be used to correct or internalize externalities was first introduced by Arthur Pigou in 1920 and has been generally accepted by economists as an efficient means to remedy inefficiencies in the allocation of resources. Pigou concluded that an environmental tax will be optimal when it is raised until the per-unit fee is just equal to the per-unit harm or “marginal social damage” (MSD) from the pollution. Intuitively this result suggests that abatement should be pursued up to the point where the marginal cost of further abatement (reflected in the emissions fee) is just equal to the marginal benefit from reducing pollution. This optimal pollution tax is widely referred to as the “Pigouvian rate.” A potential positive side effect from a policy perspective is their revenue-raising potential. “Green tax” revenues could be used in a variety of ways, such as for environmental cleanup or for research and development of clean energy technologies. But a widely discussed alternative potential use of environmental tax revenues relates to the “double dividend hypothesis,” the idea that environmental tax revenues could be used to finance reductions in preexisting taxes and, as a result, generate a second beneficial effect.

The most common motivation for taxation is, of course, to raise revenues necessary to finance the provision of public goods, such as national defense, public safety, transportation infrastructure, education, or basic research. It is recognized that these “collective consumption” or public goods would not be provided at optimal levels in private markets due to “free-riding” behavior when goods are non-rival and non-excludable. This fact underlies the rationale for government provision of public goods, and thus the need to raise revenues to finance public expenditures. Before turning fully to the double dividend question, it is important to understand several key findings from the theory of optimal taxation, or how to raise revenues needed to finance public spending in the most efficient way. A central question stems from the recognition that taxes distort behavior and, as a result, they introduce inefficiencies in the allocation of resources and, hence, a decline in social welfare compared to the (undistorted) optimum. The basic question is, how can the inefficiency, or “excess burden,” of revenue-raising taxes be kept to a minimum? Non-distortionary taxes do exist, at least in principle. A “head tax,” for example, is a lump-sum tax on individuals that would not alter incentives and behavior. A tax on “pure land rents” would be non-distortionary because land cannot be moved or increased in supply (George and Hyndman 1900). These two options, however, are not viable politically. Another possibility, in principle, is a uniform tax (a fixed percent of the price) on all goods; this would also be non-distortionary because the relative prices of all goods would remain unchanged, and thus behavior would be unaffected. Unfortunately this possibility is precluded by the presence of goods which cannot be directly taxed. Leisure is the key example. As a result of the impossibility of taxing leisure, the taxation of income or consumption will distort behavior by encouraging individuals to consume more leisure in the face of taxes on their expenditures or income. Since distortionary taxes are unavoidable, Ramsey (1927) solved the theoretical optimal tax problem which minimizes tax distortions in its basic form. The results establish that it is best to introduce a tax system that is broadly based, taxing all goods at low levels rather than a narrowly based tax system with high tax rates on a few goods. In a world

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Environmental taxation and the double dividend  253

of identical demands for all goods, equal taxes on all goods would be optimal. Where demands differ across goods but are independent of one another, the “inverse elasticity rule” holds, stating that higher taxes should be applied to goods with inelastic demands and low tax rates applied to goods where demand is elastic. It’s useful to note also that a tax on income can be thought of as being equivalent to a uniform tax on all expenditures. These theoretical findings in the economics literature rely on strong assumptions about the workings of the economy, including competitive markets, profit-maximizing firms, rational consumers, and, in mathematical terms, preferences and production relationships that conform to standard theoretical economic assumptions. Thus, it should be understood that relaxing one of these assumptions can alter the conclusions reached in this standard analysis, and thus the results and interpretations must be understood to reflect a base-case from which modifications can be, and in many cases have been, introduced and evaluated. The two main strands of tax theory just described, one to raise revenue and the other to correct externalities, were largely approached independently. Ramsey’s optimal revenue-raising taxes ignore the possibility of externalities; and Pigou’s optimal corrective tax assumes there are no revenue-raising taxes in the economy, and that the revenues generated by the environmental tax are simply returned to the economy in a “lump-sum” fashion (distributed as payments in a way that will not distort behavior). Although the integration of the two optimal tax problems was solved formally by Sandmo (1975), his conclusions received little attention for 20 years, in part because the resulting mathematical expressions are not transparent to interpret. Nevertheless, this integration of revenue-raising taxation and corrective taxation is central to issues surrounding the double dividend hypothesis.

The double dividend hypothesis

The economics of climate change drew attention to the topic of environmental taxes and the double dividend beginning in the 1990s (e.g., Pearce 1991; Bovenberg and De Mooij 1994; Bovenberg and Goulder 1996; Goodstein 2003; Manresa and Sancho 2005). While the double dividend hypothesis may

seem straightforward, the debate surrounding its validity has become complicated. This has occurred in part because views and approaches differ on how to define the double dividend hypothesis and what it implies, and there is no consensus on any single test or experiment that would validate or repudiate it. The distortions associated with a given tax or an entire tax system cannot be directly measured, and indirect estimation methods are too imprecise to provide a way to settle the question. The intuition behind the double dividend can be understood by looking at the question from two alternative starting points and asking two symmetrical questions. First, does the presence of an externality (and the opportunity to increase welfare by internalizing it with a tax) create an opportunity to lower the cost of the revenue-raising tax program in the economy generally? Second, and conversely, does the presence of public goods in need of funding (and thus a need for revenue-raising taxes) increase the benefits of an environmental tax because the revenues can be used to pay for government-provided public goods that would otherwise require revenue-raising distortionary taxes? Both thought experiments would seem to support the double dividend hypothesis. Despite this intuition, it turns out that the validity of the double dividend hypothesis depends on specific relationships in the economy (how high are preexisting tax levels? How elastic is labor supply?), and this ambiguity has made it difficult to resolve the central debate. A useful distinction is made between the “strong form” and the “weak form” of the double dividend hypothesis when applied to a standard set of assumptions about an economy (Fullerton 1997; Goulder 1995; Bovenberg 1999). The strong form of the double dividend hypothesis has been defined in several ways. The most useful of these involves a thought experiment: start from a first-best situation where an externality has been internalized with a Pigouvian tax set at the optimal level, but where there are no distortionary revenue-raising taxes (and indeed where no other externalities exist). Assume further that in this initial situation the revenue from the Pigouvian tax is just equal to the amount needed to finance government services. Then ask: If the government’s revenue requirement increases so that distortionary taxes are now necessary to raise additional William K. Jaeger

254  Elgar encyclopedia of ecological economics

revenue, will it be optimal to raise the tax on the polluting, or “dirty,” good by more, or by less, than the tax on nonpolluting, or “clean,” goods? This question was posed by Fullerton (1997), and is the same as asking whether the optimal tax differential (the difference between the tax on dirty goods and the tax on clean goods) increases or decreases as government revenue requirements increase. If the differential between the tax on the dirty good and the tax on clean goods gets larger, it means that the optimal pollution tax (differential) is getting larger, and thus has been understood to affirm the strong form of the double dividend hypothesis. The weak form of the double dividend hypothesis asks a simpler question: Does revenue recycling (the use of environmental tax revenues to finance reductions in preexisting distortionary revenue-raising taxes) produce a higher social benefit than the case where revenues were not recycled into the tax program but rather returned lump-sum to the economy (for instance, dispersed in the form of equal rebates to all households)? A simple thought experiment involving two different policies demonstrates this point intuitively. If the revenues from a Pigouvian tax were just equal to the total revenues required by government, then revenue recycling would completely eliminate the need for any distortionary revenue taxes, so there would be no distortions at all in the economy even though public goods were being provided by government. If, however, the revenues from the environmental tax were simply returned lump-sum to the economy, then distortionary revenue-raising taxes would be necessary to raise revenue for government provision of public goods, and welfare would be lowered as a result of the distortions and resulting excess burden of the tax system. Both alternatives can achieve a desired level of reduced pollution, but only with revenue recycling would the distortions from revenue-raising taxation be eliminated. There is general agreement that the weak form of the double dividend hypothesis is valid. Most of the debate surrounding the validity of the double dividend hypothesis, however, has hinged largely on the evaluation of the strong form (Fullerton 1997). However, the pivotal debate has centered on models that are “neutral,” in the sense that the demands for environmentally benign (clean) and environmentally harmful (dirty) goods are William K. Jaeger

assumed to be similar. As such, findings for these models should hold for typical or average situations.

The tax interaction effect

Beginning in the mid-1990s, the validity of the double dividend hypothesis was challenged in a set of theoretical papers that appeared to show that, despite the presence of a positive revenue-recycling effect, the optimal environmental tax would actually be lower than the Pigouvian rate when revenue-raising taxes are present (Bovenberg and De Mooij 1994; Bovenberg and Goulder 1996; Parry 1995; Fullerton 1997). These results caught many observers and even some of the authors by surprise because they seemed logically incongruous: they are at odds with the intuitive reasoning that the addition of a revenue-recycling effect would increase the benefits of green tax reform allowing the optimal environmental tax to rise above the first-best Pigouvian rate. The authors of these papers argued that their results were due to the presence of a previously unrecognized distortionary cost, which they dubbed the tax interaction effect (TIE). This effect, they argued, was negative and large enough that it would generally offset the positive revenue-recycling effect, resulting in a net welfare change lower than expected from revenue-neutral green tax reform, and as a result weakening the justification for environmental policy. Indeed, the TIE research concluded that government’s goal of providing public goods funded with tax revenue was in conflict with the goal of protecting the environment (Parry and Oates 2000). The central underlying question has been whether the welfare gains from environmental taxation in a second-best setting (where preexisting revenue-raising taxes have created unavoidable distortions and inefficiencies) are larger or smaller than in a first-best setting. In the TIE literature, however, this central question was framed indirectly by asking whether the second-best optimal environmental tax is higher or lower than the first-best Pigouvian rate; this indirect question was then tested even more indirectly by relying on a particular definition of marginal social damage (MSD) as a proxy for the Pigouvian rate. The unstated presumption here is that the value of MSD does not change when moving from a first-best to a second-best setting, so that if

Environmental taxation and the double dividend  255

the optimal pollution tax appears higher (or lower) than MSD in a second-best setting, then this means that the pollution tax has increased (or decreased).

Interpreting the tax interactions findings

With the advantage of hindsight, three factors have been found to have contributed to the negative and misleading interpretations in the TIE literature: i) the use of an unreliable benchmark, ii) an algebraic error, and iii) a failure to recognize compounding or double taxation. As a result, the apparent discovery of a large, previously unnoticed distortionary TIE appears to have been based on misleading evidence, false interpretations, or questionable assumptions (Jaeger 2011; Goodstein 2003; Schwartz and Repetto 2000). The first factor is due to the highly indirect way the TIE literature tested whether the welfare changes from environmental taxation in a second-best setting were larger or smaller than in a first-best setting. The initial logic was sound, and goes something like this: in a first-best setting, as we introduce a tax on pollution, the benefits (from internalizing the externality) outweigh the costs (the distorting effects of the tax on consumer choice) over some range. At the optimum, when the tax equals MSD, the marginal benefits are exactly equal to the marginal costs, and no further increase in the pollution tax can be justified on efficiency grounds. If, however, in a second-best setting, there is an additional benefit from revenue recycling (using the revenues to finance reductions in preexisting taxes), then the benefits from introducing and raising the environmental tax will be larger than in the first-best setting. This means that the point where the marginal benefits are just equal to the marginal costs should occur at a higher environmental tax than in the first-best case, a tax level above the first-best Pigouvian rate. Rather than carry out this test, however, the TIE literature made the test even more indirectly. Instead of comparing the value of the second-best optimal environmental tax with the value of the first-best optimal environmental tax (for instance, dollars per ton of CO­2 emissions), it was compared to MSD defined in algebraic terms rather than its numerical value at the first-best optimum.

The TIE approach then sought to test whether the second-best optimal environmental tax was higher or lower than MSD. It has been shown, however, that because MSD will vary between a first-best and a second-best setting, it is not a reliable benchmark or standard against which to compare the level of the environmental tax (Jaeger 2011; Goodstein 2003). Moreover, the correct definition of MSD is ambiguous, with three potentially valid alternatives. The numerator of MSD is the marginal social disutility from environmental damage, and this can be assumed to be constant for simplicity. The denominator of MSD is the marginal value of a unit of income and its value depends on the level of taxation. The TIE literature further defines MSD in terms of the private marginal utility of income, although the social marginal utility of income or the marginal value of public funds could also be used, and these metrics diverge from the private marginal utility of income in different ways in a second-best setting. So the test employed in the TIE literature is not reliable. The second factor reinforcing the misleading conclusions involves a straightforward algebraic error. Several authors pointed to Sandmo’s (1975) seminal optimal tax results and compared his equation for the optimal tax on dirty goods to that for the optimal tax on a clean good. They noticed that the two algebraic expressions in Sandmo’s results (one for the optimal tax on a nonpolluting good and one for the optimal tax on a polluting good) differed only by a separate term added to the expression for the tax on the dirty good. This second term was thought to be additive and equal to the environmental tax differential, and it appeared to indicate that, by inspection, this term’s value declined as taxes increased in a second-best setting – meaning that the environmental tax differential got smaller (Fullerton, 1997). This second term, however, was not simply an additive component of the overall tax – given the way the tax rate ratio was defined algebraically. TIE proponents nevertheless concluded that the Sandmo result confirmed that the pollution tax component of the optimal tax program would decline with rising revenue requirements (and with a higher marginal cost of public funds). William K. Jaeger

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A third contributing source of confusion has to do with the way revenue-raising taxes are introduced into the economy in analytical and numerical models —the “normalization” of the tax program. Revenue-raising taxes in optimal tax models are generally introduced either as a tax on wage income or as taxes on expenditures (consumption). With some simplifying assumptions, either one of these can be used to achieve the identical result: an optimal revenue-raising tax program with equal taxes on all goods or, equivalently, a single tax on wage income. Pollution taxes must necessarily be introduced on the expenditure side of the budget constraint rather than the income side: it is a tax on the dirty good. In the case where revenues are also collected with expenditure taxes, the optimal pollution tax is just added to the dirty good’s revenue-raising tax to give us a total tax on the dirty good (and the pollution component of the total tax is understood to equal the differential between the tax on clean goods and the tax on the dirty good). When revenues are collected using an income tax, however, it becomes more complicated. The presence of an income tax “compounds” the pollution tax, so that the effective tax on pollution is actually higher than its nominal value would suggest: this compounding magnifies the incentives to reduce pollution to a greater extent than is reflected in the nominal pollution tax. The pollution tax may look lower, but it still represents a higher “effective” pollution tax. Because the effective pollution tax will be higher than the nominal pollution tax in a model with income taxes, the two tax normalizations (one with income taxes and one with expenditure taxes) can represent identical optimal tax situations (in terms of incentives, allocations and welfare), but with different nominal taxes on pollution. This difference can be seen in the numerical results from an optimal carbon tax study (Bovenberg and Goulder 1996). Bovenberg and Goulder (1996) conclude that the “optimal environmental tax rates are generally below the rates suggested by the Pigovian principle – even when revenues from environmental taxes are used to cut distortionary taxes” (p. 994), and they find that this tax rate declines as revenue requirements rise. However, when their model results are re-normalized to reflect William K. Jaeger

a situation where all taxes are introduced on expenditures, the pollution tax (differential) rises with increased revenue requirements and tax levels (see Jaeger, 2011), which contradicts their interpretation. The TIE findings were controversial from the outset, but were also influential for many years. More recently, however, the pendulum appears to have swung back toward attention to revenue recycling and the double dividend hypothesis. Indeed, articles mentioning the double dividend continue to grow as do their citations, numbering in the hundreds per year in Web of Science. By contrast, a search for “tax interaction effect” finds only four articles in the past decade, and citations of this literature being less than 30 per year and declining. Unfortunately this is not a debate that can be resolved empirically. The distortionary costs of a tax cannot be directly measured or accurately estimated. The public finance literature includes numerous estimates of the distortionary costs or “marginal excess burden” of the US tax system (i.e., the distortionary cost per dollar of revenue raised to finance government) using methods that require many detailed assumptions about the responsiveness of individuals, firms, and markets, such as the labor market and capital market, to higher taxes (Auerbach and Hines 2002; Browning 1987). Large numerical computer models of the US economy have generated estimates of the distortionary costs of taxation, but the estimates can vary widely (e.g., Ballard et al. 1985; Jorgenson and Yun 1991). Similarly there are general-equilibrium simulation models that predict the impacts of the introduction of carbon tax policies in the US on carbon emissions and on the economy. Indeed, one highly regarded model among these, with many versions being revised and updated since the 1990s, finds that recycling of carbon tax revenues increases welfare (excluding the long-term climate change mitigation benefits) rather than imposing a large distortionary cost, a result that supports the strong form of the double dividend hypothesis (McFarland et al. 2018; Jorgenson et al. 2013).

Summary and conclusions

Despite a debate that has largely been played out on the basis of technical details in theo-

Environmental taxation and the double dividend  257

retical models, the intuition that an optimal revenue-raising tax program should be as broadly based as possible is also consistent with the systems approach emphasized in models of human-natural systems. This suggests that – like other goods in the economy – environmental goods and services should be priced at their social cost in a “first-best” world without distortionary taxes (the Pigouvian rate in the case of emissions). In a “second-best world,” where distortionary taxes are necessary to finance public expenditures, a broadly based revenue-raising tax should be added to all goods including environmental waste disposal services, raising their prices above their social cost – above the Pigouvian rate in the case of emissions. Despite confusion in some of the literature on this topic, the conclusions that can be drawn are in keeping with economic intuition. Environmental and revenue-raising taxes are complementary tools for achieving two kinds of government goals: the provision of public goods with revenue-motivated taxes and the protection of environmental quality with corrective taxes. Indeed, the joint pursuit of these two goals through taxation can enable government to justify doing more of each by making the optimal environmental tax higher than it would be otherwise, and by lowering the distortionary cost of financing the provision of public goods. William K. Jaeger

References

Auerbach, Alan J., and James R. Hines. 2002. “Chapter 21 Taxation and Economic Efficiency.” In Handbook of Public Economics, edited by Martin Feldstein and Alan J. Auerbach 1347–1421. Elsevier. Ballard, Charles L., John B. Shoven, and John Whalley. 1985. “General Equilibrium Computations of the Marginal Welfare Costs of Taxes in the United States.” The American Economic Review 75 (1): 128–38. Bovenberg, A. Lans. 1999. “Green Tax Reforms and the Double Dividend: An Updated Reader’s

Guide.” International Tax and Public Finance 6, 421–43. Bovenberg, A. Lans, and Ruud A. De Mooij. 1994. “Environmental Levies and Distortionary Taxation.” The American Economic Review 84 (4): 1085–9. Bovenberg, A. Lans, and Lawrence H. Goulder. 1996. “Optimal Environmental Taxation in the Presence of Other Taxes: General-Equilibrium Analyses.” The American Economic Review 86 (4): 985–1000. Browning, Edgar K. 1987. “On the Marginal Welfare Cost of Taxation.” The American Economic Review 77(1): 11–23. Fullerton, Don. 1997. “Environmental Levies and Distortionary Taxation: Comment.” The American Economic Review 87 (1): 245–51. George, Henry, and H. M. Hyndman. 1900. The Single Tax. Vierth. Goodstein, Eban. 2003. “The Death of the Pigovian Tax? Policy Implications from the Double-Dividend Debate.” Land Economics 79 (3): 402–14. Goulder, Lawrence H. 1995. “Environmental Taxation and the Double Dividend: A Reader’s Guide.” International Tax and Public Finance 2 (2): 157–83. https://​doi​.org/​10​.1007/​ BF00877495 Jaeger, W. K. 2011. “The Welfare Effects of Environmental Taxation.” Environmental and Resource Economics 49 (1): 101–19. Jorgenson, Dale W., Richard J. Goettle, Mun S. Ho, and Peter J. Wilcoxen. 2013. Double Dividend: Environmental Taxes and Fiscal Reform in the United States. MIT Press. Jorgenson, Dale W., and Kun-Young Yun. 1991. “The Excess Burden of Taxation in the United States.” Journal of Accounting, Auditing & Finance 6 (4): 487–508. Manresa, Antonio, and Ferran Sancho. 2005. “Implementing a Double Dividend: Recycling Ecotaxes towards Lower Labour Taxes.” Energy Policy 33 (12): 1577–85. McFarland, James R., Allen A. Fawcett, Adele C. Morris, John M. Reilly, and Peter J. Wincoxen. 2018. “Overview of the EMF 32 Study on U.S. Carbon Tax Scenarios.” Climate Change Economics 9 (1): 1840002. Parry, Ian W. H. 1995. “Pollution Taxes and Revenue Recycling.” Journal of Environmental Economics and Management 29 (3): S64–77. Parry, Ian W. H., and Wallace E. Oates. 2000. “Policy Analysis in the Presence of Distorting

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258  Elgar encyclopedia of ecological economics Taxes.” Journal of Policy Analysis and Management 19 (4): 603–13. Pearce, David. 1991. “The Role of Carbon Taxes in Adjusting to Global Warming.” The Economic Journal 101 (407): 938. Pigou, Arthur C. 1920. The Economics of Welfare. Macmillan & Co. Ramsey, F. P. 1927. “A Contribution to the Theory of Taxation.” The Economic Journal 37 (145):

William K. Jaeger

47. https://​doi​.org/​10​.2307/​2222721 Sandmo, Agnar. 1975. “Optimal Taxation in the Presence of Externalities.” Swedish Journal of Economics 77 (1): 86. Schwartz, Jesse, and Robert Repetto. 2000. “Nonseparable Utility and the Double Dividend Debate: Reconsidering the Tax-Interaction Effect.” Environmental and Resource Economics 15 (2): 149–57.

44. Environmentally extended multi-region input– output analysis 44.1

The need for analyzing impacts along (global) value chains

Globalization increases the interconnectedness of people and places around the world. In a connected world, goods and services consumed in one country are often produced in other countries and exchanged via international trade. Thus, consumption is increasingly met by global supply chains involving large geographical distances. As such, local consumption can lead to local and global environmental change and contribute to pollution, climate change, water, and other resource scarcity, deforestation and land conversions, and biodiversity loss, all of which affect important ecosystem services. Inequalities in consumption of goods and services are also reflected in impacts on natural resources: people in wealthier countries maintain higher incomes and more resource-intensive lifestyles, while people in poorer countries often bear the environmental consequences (Yu et al. 2013; Dorninger et al. 2021). Of course, there are huge differences in terms of income and resource consumption within countries that are also an important component of environmentally extended multi-region input–output analysis (EE-MRIO)-based analyses (e.g., Bruckner et al. 2022). Thus, consumption decisions potentially contribute to and reinforce global inequality and exploitation through global supply chains. Decision-making on environmental protection and climate change adaptation and mitigation require a better understanding of such social, economic, and environmental linkages. Therefore, we need sophisticated tools to assess a range of environmental and social implications of our consumption choices across spatial and temporal scales. To account for environmental impacts of consumption and distribution of wealth, global supply and value chain analysis is needed. Many approaches have been devel-

oped in the past to assess environmental impacts of different goods and services throughout the whole life cycle. They can be distinguished into two broad categories: bottom-up and top-down approaches. Bottom-up approaches refer to process analysis, which uses detailed descriptions of individual production processes and associated environmental impacts. However, this type of approach can lead to significant truncation errors in the calculations due to an artificial cut-off when defining the system boundaries (Feng et al. 2011). By contrast, many studies have applied top-down approaches, which rely on the regional, national, or international accounting system to show the flows of goods and services among economic sectors, and input–output analysis is one of these approaches (Hubacek and Feng 2016). EE-MRIO, based on monetary flows among sectors and regions, complemented by environmental accounts, is able to capture spatially distant links between local consumption and global environmental impacts through international trade. EE-MRIO considers the entire (global) economy as the system boundary and allows capturing environmental impacts throughout global supply chains and linking them to a wide range of final products. EE-MRIO allows linking consumption at a specific locale to the global consequences of different environmental issues, such as water and land use and stress, greenhouse gas emissions, and other air pollution, as well as economic and social indicators (Hubacek et al. 2014). One strength of the EE-MRIO framework is that it enables us to see the big picture and the interactions between different economic sectors, environmental issues, and the choices we make and how they are linked together and affect each other.

44.2

Quantification of environmental impacts associated with human consumption along entire global supply chains

To illustrate how consumption in one country can trigger pollution around the globe, we provide an example of a relatively simple product: the electric toothbrush. When a consumer purchases an electric toothbrush, that item embodies a number of production processes, each of which requires natural

259

260  Elgar encyclopedia of ecological economics Table 44.1 Environmentally extended multi-region input–output accounting framework    Country r

Country r

Country s

Final Demand r

Final Demand s

Total Output

1… j …n

1…j … n

 

 

 

​Z​ rr​

Z ​ ​ rs​

​y​ rr​

​y​ rs​

​x​ r​

​Z​ sr​

​Z​ ss​

​y​ sr​

​y​ ss​

​x​ s​

  

  

  

v​ ​ r​

v​ ​ s​   

  

  

​x​ r​

x​ ​ s​

​e​ r​

e​ ​ s​

​e​ r,dir​

e​ ​ s,dir​

1 ​⋮​ i ​⋮​ n

Country s

1 ​⋮​ i ​⋮​ n

Value Added Total Input Environmental Accounts

        

  

Note: Matrices are presented in capital letters, and vectors are presented in lower-case letters.

resources, causes emissions and other environmental impacts, but also adds jobs and triggers value added. For example, rechargeable batteries might come from Japan, the circuit board from China, and plastic parts from central Europe. All these parts are then shipped to the US where the final product is assembled and packaged. At each stage of the global value chain, pollution is produced and economic benefits are generated for countries and firms involved (Prell et al. 2014). Direct emissions associated with operating the toothbrush would accrue at the place of consumption, but indirect emissions would occur around the globe, and EE-MRIO could help us identify where those impacts happen, and who (e.g., city, country, sector, and type of household) triggers them. The EE-MRIO accounting framework consist of two parts, an MRIO table representing economic flows, and environmental accounts representing environmental flows. An MRIO table is a collection of regional input–output tables that are connected with inter-regional trade matrices. As presented in Table 44.1, an MRIO table consists of intraregional transaction blocks (i.e., Zrr and Zss) on the diagonal, and inter-regional transaction blocks (i.e., Zrs and Zsr) on the off-diagonal, intraregional (yrr and yss) and inter-regional find demand (yrs and ysr), and gross output (xr and xs), and all of these sections are in Klaus Hubacek and Kuishuang Feng

sectoral-level detail. All economic transactions in the MRIO table are in monetary values. In an MRIO framework, the production coefficient matrix A consists of sub-matrices of input coefficients for intra- and inter-regional flows. The input coefficient, ​z​  ​ ​ ​ ij​ rs​ ​​= _ ​​​x​ ​ ​ ​​, where​​ sub-matrix ​A​ rs​, is derived from A z​ij rs​ ​​is the economic flow from sector i in region r to sector j in region s, and x​ j​ s​ ​ is the total input of sector j in region s. A is a composite matrix for all regions as a whole and can be presented as: rs ij s j

rr ​ ​ rs​​  ​A  ​=  [​ ​A ​  ​  ​ ​ A ​ ​A​ sr​ ​A​ ss]​

​(44.1)​

A composite matrix of final demand and a composite vector of gross output are: ​y​ rr​ ​y​ rs​ ​Y  ​=  [​ ​  sr​  ss​] ​ ​y​  ​ ​y​  ​

​(44.2)​

r ​x  ​=  [​ ​ ​xs​  ​​ ​]​ ​x​  ​  

(​ 44.3)​

In the mathematical solution, the MRIO model can be written as: ​x   ​=    ​(​ I ​−  A)​ ​− 1​ * Y * i

(​ 44.4)​​

Environmentally extended multi-region input–output analysis  261

where i is a column vector of ones. Using L to represent the Leontief inverse matrix​   ​​(​I − A​)​ −1​​ , Equation 44.4 can be rewritten as: ​x  = ​    L * Y * i

​(44.5)​​

The environmental extension of the MRIO framework includes direct environmental impacts from all economic sectors (e.g., ​e​ r​​and​​ e​​ s​) and final consumers (e.g., e​ ​ r,dir​​and ​​e​ s,dir​​) in all regions, and environmental impacts are usually measured in physical units, such as tons, cubic meters, or hectares, based on the choice of environmental indicator. By extending the MRIO model with vectors of environmental coefficients (e.g., fs and fr), we can calculate embodied environmental impacts in inter-regional trade and consumption-based environment accounts. Environmental coefficients can be derived from _​e​x​ ​ ​ ​​​ and ​_​e​x​ ​ ​​​ for region r and region s, which indicate the environmental impacts per unit of economic output. The mathematic expression of the EE-MRIO model is shown in Equation 44.6:

tries or regions can be calculated by adding up the total embodied environmental impacts in final consumption and the household direct environmental impacts: rr ​ ​ rs​​  ​EF​​ tot​  ​=  [​ ​f​ r​   ​f​ s​  ]​*​[​L ​  ​  ​ ​ L ​*​ ​L​ sr​ ​L​ ss]​

​]​ ​+  ​[​e​ r,dir ​  ​e​ s,dir​]​ [​ ​y​  ​ sr​ ​   ​y​ ss​  ​ ​y​ rr​ y​ ​ rs​

(​ 44.8)​

where ​ ​EF​​ tot​​is a vector of environmental footprints of final consumption in different countries (r and s). The popularity of the EE-MRIO approach is not only driven by its completeness and other attractive conceptual features but also the availability of a number of global MRIO databases that have recently come online and greatly facilitated global analysis. But MRIO databases can substantially differ with regard to the number of countries they include, their sectoral detail, temporal resolution, and detail of environmental accounts. For example, the Global Trade Analysis Project database (www​.gtap​.agecon​.purdue​.edu/​) contains 65 economic sectors in each of 141 countries r ​EF​​ embodied​  ​=  [​ ​F ​ ​ ​ ​​  0s  ​​]​*​ and 19 aggregate regions in its latest version ​ ​ ​ 0 F 11, and five reference years 2004, 2007, rr rs 2011, 2014 and 2017. GTAP is one of the ​y​ rr​ ​y​ rs​ L ​   ​ ​ L ​   ​ ​ (​ 44.6)​ ​[​  sr​  ss​  ​ ​ ​*​ ​  ​  MRIO databases that has the most com​L​  ​ ​L​  ]​ [​y​ sr​ ​y​ ss]​ prehensive agricultural sectors (eight crops, where ​EF​​ embodied​is the embodied environmen- four livestock sectors, one forestry, and one tal impacts in final consumption from domes- fishing) compared with other MRIO datatic production and imports; Fr is a matrix with bases, and provides databases and models environmental coefficients on diagonal for related to greenhouse gas emissions, land use, region r and Fs is a matrix with environmental and biofuels. However, by the time of publication, the data are usually about five years coefficients on diagonal for region s. Embodied environmental impacts in old, and the latest available year is currently import of region r or export of region s can 2017 for version 11, which was released in 2022. The World Input–Output Database be calculated by: (WIOD) has 56 economic sectors (with only one agriculture sector), 43 countries, and one ​​EF​​ r, imp​ or  ​EF​​ s, exp​  ​=  F ​ ​ s​ * ​L​ sr​ *​ Rest of the World region, with environmental rr s ss sr y​ ​  ​ ​+    ​F​ ​ * ​L​  ​ * ​y​  ​  ​(44.7)​ extensions on energy use and CO2 emissions, and covers the years 2000 to 2014. The where ​ ​F​ s​ * ​L​ sr​ * ​y​ rr​​captures the embodied WIOD database focuses mainly on European environmental impacts in the import of inter- Union countries and some major economies, mediate goods and services of region r from such as China, India, Brazil, Russia, and region s to meet region r’s final demand, and​ the US), but has little detail for developing F​​ s​ * ​L​ ss​ * ​y​ sr​​captures the embodied environ- countries (wiod​.org/​home). Another widely mental impacts in the import of final goods used MRIO database is Eora (worldmrio. and services of region r from region s to meet com/). The Eora database has very good region r’s final demand. country coverage (186 countries) with a total Total consumption-based environmental of 15 909 sectors (different countries have impacts or environmental footprints of coun- different sector resolutions). In addition, it r r

s s

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provides detailed environmental information, including greenhouse gas emissions, labor inputs, air pollution, energy use, water requirements, land occupation, N and P emissions, primary inputs to agriculture (including 172 crops) from FAOSTAT, and Human Appropriation of Net Primary Productivity. It also publishes a harmonized MRIO table with 26 sectors for each country and 190 countries from 1990 to 2015. However, the harmonized table only has one agriculture sector, which is a limitation for land accounting analysis as agricultural sectors are the most land-intensive sectors. Another widely used EE-MRIO is EXIOBASE, which has 49 countries/regions, distinguishes 163 industry sectors and 200 products (eight agricultural sectors), and covers 30 types of emissions and 80 resources by industry as a monetary and .exiobase​ hybrid mixed-unit version (www​ .eu/​ ). The most recently published global MRIO with a strong focus on emerging economies is EMERGING, covering 135 sectors in 245 economies over the period 2015–19 (ceads.net). Further years are planned to be added as well as more extensive environmental accounts in addition to the currently available accounts for CO2 and material data and water consumption data. The availability of these large input– output datasets has led to scrutiny of the validity, comparability, uncertainty of the various products, and the effects of differences in aggregation. For example, Owen (2017) carried out a structural decomposition analysis to investigate the variations of regional consumption-based CO2 emissions based on three MRIO databases: Eora, GTAP, and WIOD. They found that, for a majority of regions, GTAP and WIOD tend to produce similar results. Steen-Olsen et al. (2014) found that the level of aggregation could significantly influence results depending on the purpose of the study. In general, the largest contributors to uncertainty of consumption-based emission results are – in descending order of priority – the total territorial greenhouse gas emission accounts, the allocation of emissions to economic sectors, the total and composition of final demand, and the structure of the economy. Harmonizing territorial emissions across global MRIO datasets is the single most important factor that reduces uncertainty by about 50 percent (Tukker et al. 2020). A comprehensive and systematic model found a variation of Klaus Hubacek and Kuishuang Feng

5–10 percent for both production-based and consumption-based accounts of major economies, whereas for smaller countries, variability was in the order of 20–30 percent and can reach more than 40 percent in cases of very small, highly trade-exposed countries, such as Singapore and Luxembourg (Wood et al. 2019). Thus consumption-based emission results for such countries need to be interpreted with care. More work is required to improve and institutionalize the compilation of EE-MRIO data and models to enhance the accuracy of consumption-based accounting (Tukker et al. 2018).

44.3

Outlook and conclusions

Environmental footprinting has made considerable progress over the last two decades in analyzing environmental pressures from consumption activities that arise throughout global supply chains. However, there are still a number of interesting challenges and hurdles but also encouraging developments on the way. High spatial resolution is required to model the entire global production and consumption web. This would enable the analyst to gain a better sense of the local context of consumption decisions as well as the context of resource extraction and production and the context of where impacts might occur (Hubacek et al. 2014). When focusing on the consumer end of the supply chain, we encounter the question of how to best estimate consumption patterns at a fine spatial scale considering the geographic, demographic, socio-economic, and infrastructure contexts of consumption decisions. Here “big data” in geodemographics (i.e., the development of methods for collection, integration, and analysis of massive and spatially explicit data) help with the development of small-area environmental footprint estimates (Hubacek et al. 2014). Geodemographics, based on the idea that people and places are inextricably linked, provides a new path to study human consumption patterns by embracing spatial information technologies, location-based services, social media, and other big data providers to better understand how lifestyles, infrastructure, and geographical factors shape consumption decisions. At the other end of the supply chain – that is, calculating impacts – approaches from life-cycle assessment, such as damage to human health, ecosys-

Environmentally extended multi-region input–output analysis  263

tems, and resource availability, are frequently used. This enables us, in combination with EE-MRIO, to track environmental burdens of consumption spatially and along the entire value chain. This increase in spatial granularity opens up another research frontier linking EE-MRIO with atmospheric or Earth systems science models, which can deliver more targeted environmental policies at the appropriate spatial scale as well as account for trade-offs and win–win outcomes across environmental arenas. For example, modeling trade-offs is a key component of research on the food–energy–water–climate–etc. nexus developed in response to the urgent need to enhance and integrate data and models on the interconnected energy and water challenges to inform decision-makers and the public. Consumption, on the one hand, and impacts, on the other, are linked in EE-MRIO by trade flows of goods and services, usually at the national level. These do not allow to account for differences in supply chains within countries or at great sectoral detail. Single-region input–output models also ignore regional differences but assume uniform production and consumption structures across the region under investigation. Inter-regional input–output models have been developed to enable analysis of interlinkages between regions at the subnational level. But these are often not linked to global trade models. Few studies have developed global, nested EE-MRIO models that can span multiple governance levels and potentially allow one to better fit the analysis to environmental boundaries, which rarely match administrative boundaries as provided by traditional economic datasets and models. Thus, much energy is being expended on downscaling input–output tables or upscaling based on local data to create custom-made tables fit for purpose and scale (Hubacek et al. 2014). In addition to the spatial resolution, the temporal resolution also has received considerable attention in recent years, with longer and more recent annual tables emerging; dynamics itself is barely even attempted in the EE-MRIO literature due to a high level of uncertainties in future trade flows. Nevertheless, there have been some recent advances that have built on a fairly long history of incorporating input–output models within a system dynamics approach that

provide a fruitful new direction for input– output analysis within the context of integrated assessment models (Solé et al. 2020). Last but not least, and also not to complete the list of new developments, we want to highlight the new indicators to extend the environmental and social impacts included in such analyses. For example, the “footprint family” was developed to track human pressure on the planet from different angles, such as the ecological footprint, water footprint, and carbon footprint meant to provide distinct aspects of sustainability and answer different research questions (Galli et al. 2012). While these and many other environmental indicators are complementary and cover a wide range of environmental issues, they are not well suited to deal with trade-offs that are often inherent to decision-making. Another set of footprints looks at social and economic costs or effects of environmental pollution and resource use, such as a mortality footprint (Prell et al. 2015), and numerous such social indicators have been added, including a slavery footprint, child labor footprint, or corruption footprint. While these are all very exciting new venues, they still often lack comparability and an analytic framework for evaluating the environmental, social, and economic trade-offs associated with choices. While multi-regional input–output analysis provides such a framework, its full potential has not been harnessed (Feng et al. 2019). Klaus Hubacek and Kuishuang Feng

References

Bruckner, B., K. Hubacek, S. Yuli, H. Zhong, K. Feng (2022). Impacts of poverty alleviation on national and global carbon emissions. Nature Sustainability 5, 311–20. https://​doi​.org/​10​ .1038/​s41893​-021​-00842​-z Dorninger, Christian, Alf Hornborg, David J. Abson, Henrik von Wehrden, Anke Schaffartzik, Stefan Giljum, John Oliver Engler, Robert L. Feller, Klaus Hubacek, Hanspeter Wieland (2021). Global patterns of ecologically unequal exchange: Implications for sustainability in the 21st century. Ecological Economics 179, 24. Feng, K., A. Chapagain, S. Suh, S. Pfister, K. Hubacek (2011). Comparison of bottom-up and top-down approaches to calculating the water footprints of nations. Economic Systems Research 23(4), 1–15. Feng, K., K. Hubacek, Y. Yu (2019). Local Consumption and Global Environmental

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264  Elgar encyclopedia of ecological economics Impacts: Accounting, Trade-Offs and Sustainability. Routledge. Galli, A., T. O. Wiedmann, E. Ercin, D. Knoblauch, B. R. Ewing, S. Giljum (2012). Integrating ecological, carbon and water footprint into a “footprint family” of indicators: Definition and role in tracking human pressure on the planet. Ecological Indicators 16, 100–12. https://​ doi​ .org/​10​.1016/​j​.ecolind​.2011​.06​.017 Hubacek, K., K. Feng (2016). Comparing apples and oranges: Some confusion about using and interpreting physical trade matrices versus multi-regional input–output analysis. Land Use Policy 50, 194–201. Hubacek, K., K. Feng, J. Minx, S. Pfister, N. Zhou (2014). Teleconnecting consumption to environmental impacts at multiple spatial scales – Research frontiers in environmental footprinting. Industrial Ecology 18(1), 7–9. Owen, A. (2017). Techniques for Evaluating the Differences in Multiregional Input–Output Databases: A Comparative Evaluation of CO2 Consumption-Based Accounts Calculated Using Eora, GTAP and WIOD. Springer International Publishing. Prell, C., K. Feng, L. Sun, M. Geores, K. Hubacek (2014). The global economic gains and environmental losses of US consumption: A world-systems and input–output approach. Social Forces 93(1), 405–28. Prell, C., L. Sun, K. Feng, T. W. Myroniuk (2015). Inequalities in global trade: A cross-country comparison of trade network position, eco-

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nomic wealth, pollution and mortality. PLoS ONE 10(12), e0144453. Solé, J., R. Samsó, E. García-Ladona, A. García-Olivares, J. Ballabrera-Poy, T. Madurell, A. Turiel, et al. (2020). Modelling the renewable transition: Scenarios and pathways for a decarbonized future using Pymedeas, a new open-source energy systems model. Renewable and Sustainable Energy Reviews 132, 110105. Steen-Olsen, Kjartan, Anne Owen, Edgar G. Hertwich, Manfred Lenzen (2014). Effects of sector aggregation on CO2 multipliers in multiregional input–output analyses. Economic Systems Research 26(3), 284–302. Tukker, Arnold, Arjan de Koning, Anne Owen, Stephan Lutter, Martin Bruckner, Stefan Giljum, Konstantin Stadler, Richard Wood, Rutger Hoekstra (2018). Towards robust, authoritative assessments of environmental impacts embodied in trade: Current state and recommendations. Journal of Industrial Ecology, 22, 585–98. https://​doi​.org/​10​.1111/​ jiec​.12716 Tukker, Arnold, Richard Wood, Sarah Schmidt (2020). Towards accepted procedures for calculating international consumption-based carbon accounts. Climate Policy 20(1), S90–S106. Wood, R., D. D. Moran, J. F. D. Rodrigues, K. Stadler (2019). Variation in trends of consumption based carbon accounts. Scientific Data 6(1), 99. https://​doi​.org/​10​.1038/​s41597​-019​ -0102​-x Yu, Y., K. Hubacek, K. Feng (2013). Tele-connecting local consumption to global land use. Global Environmental Change 23(5), 1178–86.

45. Ethics of quantification The topic

We describe here the ethics of quantification and provide recommended readings on the topic. When talking about quantification – dividing and enumerating the real in discrete quanta, one should be aware of the many existing families where production of numbers takes place from modelling to statistical indicators and statistical inference, from algorithms to rating and ranking. As noted by Mennicken and Espeland (2019), caution should be taken “against unifying accounts of quantification” – minding instead “the importance of tracking quantification across different sites” (223). At the same time, Popp Berman and Hirschman (2018) ask, “what qualities are specific to rankings, or indicators, or models, or algorithms?” (265). Quantifications are thus different and yet share important attributes, as also demonstrated by the sprouting of neologisms such as “numerification” and “datafication” applied to different instances of quantification. We highlight here the common ingredients while being mindful of the differences. Quantification is a key element of our history and of our modernity. Descartes opened a longstanding conceptual paradigm of using “geometry as the model for true and certain knowledge” (Ravetz, cited in Pereira & Funtowicz 2015). The Cartesian dream of mastering and possessing nature inevitably includes a quantification agenda. Later, Condorcet embraced this quantitative understanding, even to adjudicate social choices (Pereira & Funtowicz 2015), by developing – or rather rediscovering1 – one of the algorithms for what we would call today multi-criteria analysis (Feldman 2005). For Galileo, the book of nature is written in the language of mathematics. The historical perspective tells us that techniques of quantification and visualization allowed in the West an extraordinary acceleration. Started in the 14th century, this acceleration involved astronomy, music, painting, navigation, ballistics, cartography, and many other arts and crafts, eventually leading to the domination of Europe over the rest of the world (Crosby 1996).

Nowadays, our society is imbued with different families of quantification – in statistics, modelling, data science, and the movement for a quantified self, with a remarkable degree of contamination or cross-fertilization among them (Mennicken & Espeland 2019). Using the full spectrum of tools, from statistical inference to mathematical modelling to machine learning, is also crucial for collecting, producing, and advancing scientific knowledge. Those tools are essential for most disciplinary fields of research, including ecology. The field of ecology has a rich and fruitful relation to activities of quantification, including Robert Rosen’s work on mathematical modelling of natural systems and the definition of anticipatory systems as systems possessing a model of themselves that are, thus, capable of orienting their reaction to the environment (Louie 2010; Rosen 1991). At the same time, the fathers of the ecological movement have always been circumspect in the use of quantification, calling for careful application. For example, Fritz Schumacher (1973) warns that, while quantitative differences are easier to communicate than qualitative differences, their immediateness and precision may result from the suppression of crucial differences of quality. The epistemological movement of post-normal science (PNS) reacted to the excess of certainty brought about by the methodology of quantification used in environmental assessment (Funtowicz & Ravetz 1993). A very famous PNS paper titled “The Value of a Songbird” challenged at the time the prevailing notion of valuing environmental goods in monetary terms (Funtowicz & Ravetz 1994). Scholars of science and technology studies warn that quantification often accompanies reductionism, the reformulation of a political problem into a technical one (Ravetz 1971; Scoones & Stirling 2020). This considerable debate bears a clear message: The ideal of rationality and universalism conveyed by numbers should be scrutinized. Several implications rising out of a simple number need to be detected and clearly addressed. One should consider the dominating role that numbers play in determining principles and values, such as democracy, merit, participation, accountability, and even “fairness”. For all these, there are evidently other dimensions at play that numbers cannot capture, or distortion of the real resulting from numbers

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taking centre stage. In this way, existing numbers decide how society counts and how it will count in the future. They are “reactive” (i.e. they change what they touch; Espeland & Sauder 2016); a vivid example is higher education, transformed into a global market by the international rankings (Éloire 2010). Once put into action, the “inertia of data” (Merry 2016) can strengthen the role of quantification and create a dependency: suffice to think of gross domestic product, still reigning as the uncontested measure of well-being in spite of a rich production of alternative measures produced by the movement for sustainability (Boulanger 2018). For data scientist Cathy O’Neil (2016), numbers in the form of ratings and algorithms may generate “pernicious feedback loop[s]” and “create the environment that justifies their assumptions” (29). While the seduction of numbers (Merry 2016) is generally associated with their apparent neutrality, numbers play important socio-political roles. They provide to their makers epistemic legitimacy – which may translate into power (Porter 1995) and, in extreme cases, may hide arbitrary constructs and practices, as was the case for the financial products leading the way to the latest recession (Porter 2012). Porter (1995) notes how the quest for objectivity conferred by numbers is often defensive, and deployed in cases where the elites are weak, trust is eroded, and private interests dominate. For an ecologist, Porter offers an instructive discussion of how cost–benefit analysis – an apparent paradigm of impersonal and dispassionate analysis – was in fact the theatre of inter-agency fights in the US federal bureaucracy, where each side defended its brand of objectivity upheld by specific strategies of computation. Ecologist Langdon Winner (1989) discusses risk analyses and cost– benefit analyses critically: Why should one accept a new product or practice because it is “safe” when the introduction of the same practice was not negotiated beforehand with stakeholders? Measures of risk have been seen as ethically non-neutral (Saltelli et al. 2020b) in the disputes over genetically modified products. For some, a focus on risk measures may distract from questions about the desirability of the same product or about the power relations engendered by their introduction (Marris 2001). Andrea Saltelli and Monica Di Fiore

Anchoring a political debate on the wrong number or set of numbers may limit the space of the available policy options and lead to poor choices. Foremost numbers should not be produced in order to offer politicians the opportunity to abdicate decisions by transforming a political decision into a technical one. Colonizing private and public citizens’ lives and democratic deliberation is among the many alarms about the extraordinary development of algorithms and big data. For Shoshana Zuboff (2019), the owners of major platforms (such as Facebook and Amazon) have developed a novel “instrumentarian power” that allows them to orient the behaviour of consumers and voters. For the jurist Alain Supiot (2007), a system of number-based rules is taking the place of governing by just laws. His book, Governing by Numbers, warns that governing by algorithm empties both democracy and the rule of law, with danger of a re-feudalization of society. For the French statactivistes, an important current of sociology of numbers that takes inspiration from thinkers such as Alain Desrosieres, Pierre Bourdieu, and the same Bruno Latour, “another number is possible”, meaning that wrong measures (e.g., wrong indicators) can be fought with better ones. Several academic or grassroots initiatives have sprouted to contrast numerification and datafication, such as the Algorithmic Justice League (2020) in the US, and the Data Justice Lab (Cardiff University 2020) and the Radical Statistics Group (2020) in the UK. In France, an international project of sociology of quantification is led by the CNRS (French National Research Institute for Sustainable Development 2020). In this emerging scenario, some scholars have called for an ethics of quantification (Espeland & Stevens 2008), arguing that numbers shape the real in a way that can serve humankind, as well as distort it, by privileging the measurable at the expense of the unmeasurable. One should instead scrutinize any conceit that measurement provides privileged or exclusive access to the real. One important rationale for an ethics of quantification is the non-appealability of decisions reached algorithmically – one cannot complain about being denied a mortgage, as the same bank officer cannot look inside the proprietary software that reached the decision. Additionally, the recourse to

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quantification may evade a process of negotiation and deliberation (e.g., when software is used to adjudicate disputes in tribunals or administrations; Brauneis & Goodman 2018). Ethical concerns are also present in the quantification ruled by statistics. True statistical wars are taking place around the subject of statistical inference, whose misuse is held partly responsible for an ongoing reproducibility crisis affecting science (Leek et al. 2017; Saltelli & Funtowicz 2017). The situation is possibly even worse in the field of mathematical modelling, where the lack of a disciplinary house permits an incredible diaspora of methods of validation and verification (Padilla et al. 2018). Here is where an ecologist can find inspiration for their daily chores for an ethical or responsible quantification. Targeting our remarks to the different contexts of quantification, we note: ● In the context of hypothesis testing, many non-statistician (and some statisticians, apparently) misuse the p-test. For our purpose, suffice it to note that passing a test at the 5 per cent significance level does not imply almost-sure success; it does not correspond to a 1-in-20 (5 over 100) chance of being wrong. This chance – the so-called false discovery rate – can be much higher, and close to one in three, even under “good” operating conditions (Colquhoun 2014). ● In the context of modelling, as noted in a recent manifesto, there are golden rules to be followed for responsible modelling, which touch upon the need for rigorous uncertainty and sensitivity analysis, self-restraint in the complexity of models to avoid modelling hubris, attention to hidden or forgotten assumptions, and candour with respect to deep uncertainty or outright ignorance (Saltelli et al. 2020a). ● Sensitivity auditing – an extension of sensitivity analysis to the normative aspects of modelling, can find application to several families of quantification (Saltelli et al. 2013). ● The idea of coproduction – generating numbers with the subjects measured or affected, permeates different families of quantification, from statistics (Salais 2022) to modelling (Ravetz 2003) to algorithms (Amoore 2020).

In any instance, ecologists should keep in mind that, at any of the many steps leading to a quantification, uncertainty and ambiguity may lurk in the penumbra – a phenomenon that never ceases to surprise the same experts (Breznau et al. 2021). It is worth noting that issues of quantification enter the replicability crisis (Saltelli & Funtowicz 2017) via two processes – on the one hand, the system of metrics that is predominating in the appraisal of academic merits is partly responsible for the culture of publish or perish that affects academic production (Ioannidis et al. 2014; Smaldino & McElreath 2016). On the other hand, this culture is also partly responsible for the misuse of statistical methods when corners are cut in order to achieve a published result. A mission to be entrusted to the ethics of quantification is surely to contrast the prevailing narrative in which numbers provide a reassuring role of neutrality (numbers as fact) and control (all can be computed). An ecologist should be alert to the non-neutrality of numbers and to the same methods of quantification: a cost–benefit analysis is unlikely to yield the same result as a multi-criteria analysis (or a social multi-criteria analysis; Saltelli et al. 2020b). Numbers in a culture of hubris should be replaced by numbers in a culture of humility (Jasanoff 2003). A recent review covering many of the themes of the present entry is Mennicken and Espeland (2019), while the reasons for an ethics of quantification are further discussed in Saltelli et al. (2021) and Di Fiore et al. (2023). Implications for policy are discussed in Saltelli and Di Fiore (2020), who debate how scientific knowledge translates into political knowledge and, then, into a policy decision. A paradox noted by sociologists of quantification is that scientific knowledge may narrow, rather than expand, the space of policy options, as well as the choice of the tools deployed for their implementation. Thus, science may contribute to a “closing down” of the uncertainties (Scoones & Stirling 2020) via processes that are variously noted in the literature as reductionism, depoliticization, or technocratic orthodoxy (van Zwanenberg 2020). A paradox noted in Saltelli and Di Fiore (2020) is that, while practices of quantification in decision-making obey criteria of utility and efficiency, their impact can strengthen lock-ins and path Andrea Saltelli and Monica Di Fiore

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dependency, and ultimately undermine their logic of efficiency. Andrea Saltelli and Monica Di Fiore

Development. (2020). SSSQ – Society for the Social Studies of Quantification. https://​ en​.ird​.fr/​project​-sssq​-society​-social​-studies​ -quantification Funtowicz, S., & Ravetz, J. R. (1993). Science for the post-normal age. Futures, 25(7), 739–55. Note Funtowicz, S., & Ravetz, J. R. (1994). The 1. The method of Condorcet was first devised by the worth of a songbird: Ecological economics as Catalan thinker Ramon Llull circa 1299 (Munda a post-normal science. Ecological Economics, 2008). 10(3), 197–207. Ioannidis, J. P. A., Forstman, B., Boutron, I., Yu, References L., & Cook, J. (2014). How to make more published research true. PLoS Medicine, 11(10), Algorithmic Justice League. (2020). Algorithmic e1001747. Justice League – Unmasking AI harms and Jasanoff, S. (2003). Technologies of humility: biases. https://​www​.ajl​.org/​ Citizen participation in governing science. Amoore, L. (2020). Cloud Ethics, Algorithms and Minerva, 41, 223–44. the Attributes of Ourselves and Others, Duke Leek, J., McShane, B. B., Gelman, A., Colquhoun, University Press. D., Nuijten, M. B., & Goodman, S. N. (2017). Boulanger, P. M. (2018). A systems-theoretical Five ways to fix statistics. Nature, 551, 557–9. perspective on sustainable development and indicators. In S. Bell & S. Morse, eds., Routledge Louie, A. H. (2010). Robert Rosen’s anticipatory systems. Foresight, 12(3), 18–29. Handbook of Sustainability Indicators and Marris, C. (2001). Final Report of the PABE Indices, 124–41. Routledge. research project funded by the Commission Brauneis, R., & Goodman, E. P. (2018). of European Communities, Contract number: Algorithmic transparency for the smart city. FAIR CT98-3844 (DG 12 - SSMI). Yale Journal of Law & Technology, 20, 103–76. Breznau, N., Rinke, E. M., Wuttke, A., . . . Mennicken, A., & Espeland, W. N. (2019). What’s new with numbers? Sociological approaches to Nguyen, H. H. V. (2021, March 24). Observing the study of quantification. Annual Review of many researchers using the same data and Sociology, 45(1), 223–45. hypothesis reveals a hidden universe of uncertainty. MetaArXiv. https://​doi​.org/​10​.1073/​ Merry, E. S. (2016). The Seductions of Quantification: Measuring Human Rights, pnas​.2203150119 Gender Violence, and Sex Trafficking. Cardiff University. (2020). Data Justice Lab. University of Chicago Press. https://​datajusticelab​.org/​ Colquhoun, D. (2014). An investigation of the Munda, G. (2008). Social Multi-Criteria Evaluation for a Sustainable Economy. false discovery rate and the misinterpretation Springer. of p-values. Royal Society Open Science, 1, O’Neil, C. (2016). Weapons of Math Destruction: 140216. How Big Data Increases Inequality and Crosby, A. W. (1996). The Measure of Reality: Threatens Democracy, Random House Quantification in Western Europe, 1250–1600, Publishing Group. Cambridge University Press. https://​doi​.org/​10​ Padilla, J. J., Diallo, S. Y., Lynch, C. J., & Gore, .1017/​CBO9781107050518 R. (2018). Observations on the practice and proDi Fiore, M., Kuc‑Czarnecka, M., Lo Piano, S., fession of modeling and simulation: A survey Puy, A. and Saltelli, A. (2022). The Challenge approach. SIMULATION, 94(6), 493–506. of Quantification: An Interdisciplinary Reading, Minerva, 61, 53–70 (2023). https://​doi​.org/​10​ Pereira, Â. G., & Funtowicz, S. (2015). Science, Philosophy and Sustainability: The End of the .1007 Cartesian Dream. Routledge. Éloire, F. (2010). Le classement de Shanghai. Histoire, analyse et critique. L’Homme et La Popp Berman, E., & Hirschman, D. (2018). The sociology of quantification: Where are we now? Société, 178(4), 17. Contemporary Sociology, 47(3), 257–66. Espeland, W. N., & Sauder, M. (2016). Engines of Anxiety: Academic Rankings, Reputation, and Porter, T. M. (1995). Trust in Numbers: The Pursuit of Objectivity in Science and Public Accountability, Russell Sage Foundation. Life. Princeton University Press. Espeland, W. N., & Stevens, M. L. (2008). A sociology of quantification. European Journal of Porter, T. M. (2012). Funny numbers. Culture Unbound, 4, 585–98. Sociology, 49(3), 401–36. Feldman, J. (2005). Condorcet et la mathéma- Radical Statistics Group. (2020). Radical Statistics. https://​www​.radstats​.org​.uk/​ tique sociale. Enthousiasmes et bémols. Mathématiques et Sciences Humaines, 172(4), Ravetz, J. R. (1971). Scientific Knowledge and Its Social Problems. Oxford University Press. 7–41. French National Research Institute for Sustainable Ravetz, J. R. (2003). Models as metaphors. In B. Kasemir, J. Jager, C. Jaeger, M. Gardner, eds.,

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Saltelli, A., & Funtowicz, S. (2017). What is science’s crisis really about? Futures, 91, 5–11. Saltelli, A., Guimaraes Pereira, Â., van der Sluijs, J. P., & Funtowicz, S. (2013). What do I make of your latinorum? Sensitivity auditing of mathematical modelling. International Journal of Foresight and Innovation Policy, 9(2–4), 213–34. Schumacher, E. F. (1973). Small is Beautiful: Economics as if People Mattered. Harper Perennial. Scoones, I., & Stirling, A., eds. (2020). The Politics of Uncertainty. Routledge. Smaldino, P. E., & McElreath, R. (2016). The natural selection of bad science. Royal Society Open Science, 3, 160384. Supiot, A. (2007). Governance by Numbers: The Making of a Legal Model of Allegiance. Oxford University Press. van Zwanenberg, P. (2020). The unravelling of technocratic orthodoxy. In I. Scoones & A. Stirling, eds., The Politics of Uncertainty, 58–72. Routledge. Winner, Langdon. (1989). The Whale and the Reactor: A Search for Limits in an Age of High Technology. University of Chicago Press. Zuboff, S. (2019). The Age of Surveillance Capitalism: The Fight for a Human Future at the New Frontier of Power. PublicAffairs.

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46. Fetish, commodity fetishism and ecosystem services The term “fetish” was originally coined by 16th-century Portuguese colonialists to refer to the mystical human properties of objects – in this paper we will not use the term fetish following this sense, in the Freudian sense as humanising objects. Later, during and after the Industrial Revolution, the concept of fetishism was used to refer to reification or, in other words, to the process of objectifying human beings and human relations. In turn, “commodity fetishism” can be traced back to the work of Karl Marx (1867) who, in the first volume of Das Kapital from 1867, defined it as “nothing but the definite social relation between men themselves which assumes here, for them, the fantastic form of a relation between things . . . the products of the human brain appear as autonomous figures endowed with a life of their own, which enter into relations both with each other and with the human race”. The term, however, has evolved to include other scholarly perspectives, and in particular to go beyond the domain of industrial societies. Generally speaking, Nelson (2001) defines it as the means by which “a variety of distinctive incomparable goods and services can be made comparable by a market exchange” (503). Hornborg (2001) notes that anthropology of religion refers to it as the “attribution of human mental faculties to non-human objects” (474). Researchers have also typified commodity fetishism on the nature of economic transactions. Gift economies personify objects and create qualitative relations between them, while commodity economies treat human parts as objects and are meant to establish quantitative equivalence value between objects (Graeber 2001; see Box 46.1 for three accounts of commodity fetishism and biodiversity in market and gift economies). Some Global Value Chain (GVC) researchers examine fetishism in the context of the politics of consumption, while others link it to the availability of information on the social and environmental conditions of production of any particular commodity (see Bernstein and Campling 2006). GVC analyses help identify key issues regarding scale and com-

petition in the capitalist trading system, and one can assume from them that solutions to the invisibility of labour and natural capital embedded in commodities may arise from upgrading actors in the value chain, although there would always be some actors downgraded and others upgraded in this chain. In other words, trading will always require unequal power asymmetries in the production, distribution, and consumption process. Other scholars take a cultural perspective on commodity fetishism, including those referring to shopping as a social practice and the principal means of shaping an identity (Appadurai 1986). An illustrative analysis of this approach alludes to fair trade and exchange relations in coffee trading (Hudson and Hudson 2003). These authors suggest that the “alternative-trade movement represent an initial attempt to counter the pervasiveness of commodity fetishism” (Hudson and Hudson 2003, 413) because access to information attempts to unveil the social and natural relationships within the process of production. Commodity fetishism is thus understood as the masking of the social relationships underlying the process of production. Hudson and Hudson (2003) also acknowledge that fair trade does not reach a broad segment of consumers and that labelling lacks transparency, but they fail to recognise that commodity fetishism can also be reinforced through the price system governing fair trade and other commodity markets. In the context of marketing environmental services (ES), commodity fetishism becomes a useful analytical concept because it allows us to define ES as commodities, relates use values to exchange values, and links supply and demand for these services, thus unveiling the natural and social relations embedded in the production and exchange processes.

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BOX 46.1 BIODIVERSITY COMMODIFIED The first step in the commoditisation process is to define a commodity, which for biodiversity means the selection of a single measurement from a single discipline that is instrumental for the biodiversity-commodity market. Biodiversity characterisation depends on three main disciplines: taxonomy, genetics, and ecology (Bisby 1995). The different areas of study aim at providing a ref-

Fetish, commodity fetishism and ecosystem services  271 erence system for all organisms, knowledge on the gene variations within and between species, and knowledge on the varied ecological systems in which taxonomic and genetic diversity is located, respectively (Chopra and Kumar 2004). However, it is not possible to set up a measurement system that describes all the components of biodiversity from the different disciplines and within a single discipline; this methodological effort is also helpless, as different measurements of biodiversity depict a limited view of life on Earth. In 1991, for instance, INBio (the Costa Rican institute for biodiversity) and Merck (the powerful pharmaceutical company for the US) signed an agreement for prospecting forest diversity with the aim of finding pharmaceutical active ingredients. The main definition of biodiversity employed for this commoditisation of biodiversity was genetic diversity, particularly genes. This definition enabled Merck full access to the Costa Rican forest’s genetic diversity, including useful enzymatic catalytic precursors. In other words, biodiversity, which is the variety and variability of different life forms, is narrowed down to a compendium of genes useful for pharmaceutical interests under the assumption that such bioprospecting-related funding will help preserve the rest of the inherent complexity. The second step in the commoditisation process is assigning a unique exchange value to an already defined commodity. Biodiversity has been an object of exchange in many different cultures for a very long time. Different types of markets have been concerned with these exchanges. For instance, in the case of Melanesians, wives, like husbands, help raise pigs to show their commitment to their marriage (original relation). The pig is an embodiment of that relation until it leaves the domestic sphere and enters the public sphere of male ceremonial exchange, where its value shifts, and it comes to embody the importance of relations between men (new relation). (Graeber, 2001, 41, emphasis and comments added) To assign an exchange value means to prioritise and impose a worldview above others. In the case of biodiversity, if Melanesian pig exchange traditionally happened through gifts, and now the dominant trading system is

a monetary market, then a whole worldview is at stake. Humans and other species have been coevolving for millennia; it can also be said that the multiplicity of trading systems is an expression of a co-evolutionary system, which is represented in the multiplicity of worldviews of humankind. The third step towards the commoditisation of biodiversity is the creation of standard units of measurements. The Service Providing Unit (SPU) was developed as a concept linking a group of individuals and the ecosystem service it provides, and is then used as a criterion to delineate populations (Luck et al. 2003). In Peru, potatoes have been traded for millennia; there are 247 identified varieties of potatoes subjected to local trading outside monetary markets. They have mainly been exchanged through bartering in rural markets (chalayplassa; International Institute for Environment and Development [IIED], 2005), and low-land peasants and high-land peasants barter their crops using socially agreed measurements. Farming and growing potatoes for trading in monetary markets required the development of a few stable varieties of potatoes (SPU of commodified potatoes). Ecological concepts are now subsumed to commoditised-ecological concepts, which are then denuded of any biological and evolutionary background and directly respond to monetary market forces. Market potato varieties no longer have a biological meaning and, as an SPU, simply respond to monetary market forces; they hardly describe scarcity or abundance of potatoes in the wild.

Nicolas Kosoy

References

Appadurai, A., 1986. The Social Life of Things: Commodities in a Cultural Perspective. Cambridge University Press, Cambridge. Bernstein, H., Campling, L., 2006. Commodity studies and commodity fetishism II: “profits with principles”? Journal of Agrarian Change, 6(3): 414–47. Bisby, F.A., 1995. Characterization of biodiversity. In Heywood, V.H. (Ed.), Global Biodiversity Assessment, 21–106. Cambridge University Press, Cambridge. Chopra, K., Kumar, P., 2004. Forest biodiversity and timber extraction: an analysis of the interaction of market and non-market mechanisms. Ecological Economics, 49: 135–48. Graeber, D., 2001. Toward an Anthropological Theory of Value. Palgrave, New York. Hornborg, A., 1998. Towards an ecological theory of unequal exchange: articulating world system

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272  Elgar encyclopedia of ecological economics theory and ecological economics. Ecological Economics, 25: 127–36. Hornborg, A. 2001. Symbolic technologies. Machines and the Marxian notion of fetishism. Anthropological Theory 1(4): 473–496. doi:10.1177/14634990122228854 Hudson, I., Hudson, M., 2003. Removing the veil? Commodity fetishism, fair trade, and the environment. Organisation & Environment, 16(10): 413–30. International Institute for Environment and Development (IIED), 2005. Traditional Resources, Rights and Indigenous Peoples in

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the Andes. IIED Reclaiming Diversity Series, IIED, London. Luck, G.W., Daily, G.C., Ehrlich, P.R., 2003. Population diversity and ecosystem services. TRENDS in Ecology and Evolution, 18(7): 331–6. Marx, K., 1867. Capital, Volume 1, Part 1: Commodities and Money. http://​www​.marxists​ .org/​archive/​marx/​works/​1867​-c1/​ch01​.htm Nelson, A., 2001. The poverty of money: Marxian insights for ecological economists. Ecological Economics, 36: 499–511.

47. Future generations Concepts of intergenerational justice have played a central role in ecological economics since the field was established in the late 1980s. Building on the seminal contributions of Boulding (1966) and Daly (1977), ecological economists converged on the importance of transitioning from what Boulding termed a “Cowboy Economy,” dependent on nonrenewable resources, to a “Spaceman Economy” in which materials were recycled while energy was derived from sunlight and the wind. Setting aside questions of feasibility and timing, this entry focuses on how the need for a sustainability transition is based on the belief that ecological-economic systems should be governed in a way that secures the rights and interests of future generations in the long run. From an ethical perspective, the moral status of future generations is more delicate than might appear at first. From the perspective of Classical Utilitarianism, equal weight should be attached to the well-being of every present and future person. In the context of climate stabilization, this premise implies that urgent steps should be taken to rapidly decarbonize the world economy (Cline, 1992; Stern, 2007). This follows from the fact that the long-run benefits of climate stabilization – measured in terms of contributions to well-being – are much larger than the short-run costs of decarbonization. Attaching equal weight to long-term benefits thus tips the balance toward aggressive action as a matter of moral priority. This argument, however, runs into a number of possible objections. First, utilitarians argue that transferring resources from the relatively wealthy to the relatively poor typically leads to net improvements in overall social welfare (Singer, 2002). Yet the observed levels of inequality both within and between nations and world regions suggest that decision-makers – including elected democratic governments – typically do not act in the manner prescribed by utilitarianism. Second, when applied across intergenerational timescales, utilitarianism would seem to mandate sharp reductions in current consumption levels to invest more heavily in the future. That’s true even though future generations will likely enjoy a higher standard of living than present persons – and especially

those living in poverty – without taking such drastic steps. Broome (2008) aims to resolve this paradox by arguing that extra weight should be assigned to the welfare of the worst-off members of society (both within and between generations). Drawing on the work of Koopmans (1972), Dasgupta and Heal (1979) present an axiomatic framework in which the utility or welfare of future generations should be discounted relative to the present. Under plausible assumptions concerning the relationship between material affluence and well-being, Nordhaus (2007) argues that a utility discount rate of about 1.5 percent per year is most consistent with observed rates of capital investment and economic growth. This implies that one unit of welfare accruing 100 years from today has a present value of just 0.23 units. On the one hand, this may be consistent with an optimal path in which consumption and welfare grow substantially over time, leaving future generations better off than people living today. On the other hand, it suggests that relatively modest steps toward climate stabilization might be economically optimal, with the costs of climate change offset by improvements in the standard of living. The practice of utility discounting has long been controversial. Ramsey (1928), for example, argued that it is “ethically indefensible and arises merely from the weakness of the imagination” (543). Howarth (2011) notes that this approach might accurately reflect the degree of altruism that present society holds toward posterity. But neither utilitarian nor deontological moral theories ground a person’s moral standing on the degree of altruism held by others. On the contrary, moral responsibilities sometimes oblige us to take actions that go beyond the scope of our subjective desire to benefit others. A final issue with utilitarianism relates to the endogeneity of human population over intergenerational timescales (Sidgwick, 1901, Book IV, ch. 1). The question is whether it is better to create a future in which: (A) a large number of people have a relatively low quality of life; or (B) a small number of people would enjoy opulence and prosperity. Under Total Utilitarianism, Option (A) should be chosen if it maximizes the sum total of welfare across all persons over time. In contrast, Average Utilitarianism – which aims to maximize the well-being of a typical person – would favor Option (B). It is unclear that either of these

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approaches fits with most people’s intuitions concerning duties to future generations. And yet the utilitarian impulse that consequences should matter in evaluating decision options does have important resonance. The other main approach to questions of intergenerational justice builds on principles from deontological or rights-based ethics. This is seen in the Brundtland Commission’s famous definition of “sustainable development,” which entails meeting present needs “without compromising the ability of future generations to meet their own needs” (World Commission on Environment and Development, 1987, 42). In this framework, the present generation has an obligation to provide future generations with a “structured bequest package” (Norton and Toman, 1997) that secures the ability of future persons to define and pursue their own conception of the good life. This framework can be grounded on Sen’s (2009) Capabilities Approach, which provides one pillar for Raworth’s (2017) so-called “Doughnut Economy.” Raworth specifies a long list of capabilities that provide the foundation for a “safe and just space for humanity,” including (but not limited to) basic needs for food, water, and energy along with social needs such as education, gender equality, and political voice. Raworth integrates this with the “Planetary Boundaries” approach of Rockström et al. (2009), who emphasize the need to stabilize the major Earth systems that support and sustain social and economic activity, including global climate, the ozone layer, biodiversity, and nutrient cycling. These recent contributions operationalize the concept of “strong sustainability,” in which future generations are seen as having an ex ante right to inherit an undiminished set of environmental resources (Ekins et al., 2003). While this approach is intuitively appealing, it presents a very important philosophical question: In what sense can future generations be said to have well-defined rights? On this point, Parfit (1984) expresses a high degree of skepticism. Consider a choice between two development pathways, one involving a low standard of living and a badly degraded natural environment, the other involving a utopian future characterized by prosperity, social justice, and sustainability. Parfit’s argument is that – due to the minute details of reproductive decisions – different sets of potential persons would be born into these Richard B. Howarth

two worlds. If this were correct, then choosing the sustainable future would not improve the welfare of persons who otherwise would endure a low quality of life. On the contrary, the people born into the degraded future might have reason to thank us for choosing to bring them into being, as long as they had lives that were at least minimally worth living. Parfit’s argument is clever and makes an important point that resonates with the debate over Total vs. Average Utilitarianism. Yet it overlooks one very important point that is central to questions of intergenerational justice. In particular, today’s adults have direct obligations to their living children and grandchildren. Part of what we owe to them is an ability to raise their own families without impoverishing themselves. That is to say, the ability to have and care for children is itself a centrally important capability in Sen’s sense. If we agree that it would be wrong for the present generation to act in ways that – although self-beneficial – imposed harms and hardships on their existing children and grandchildren, it follows that a “chain of obligation” exists that connects the present with the very long run, thereby dissolving Parfit’s Paradox (Howarth, 1992). In contrast with the Capabilities Approach and the concept of “strong sustainability,” advocates of “weak sustainability” argue that obligations to future generations are more limited. Solow (1993), for example, asserts that: The duty imposed by sustainability is to bequeath to posterity not any particular thing . . . but rather to endow them with whatever it takes to achieve a standard of living at least as good as our own and to look after their next generation similarly. We are not to consume humanity’s capital, in the broadest sense. (168)

Solow’s point is in one sense well taken. Surely, a plausible theory of intergenerational justice would focus on the importance of maintaining or improving human well-being from each generation to the next. And in some contexts, the depletion of natural resources might be offset by corresponding investments in new technologies and capital equipment. This is seen, for example, in the transition from a fossil-fuel-based economy to an economy sustained by high energy efficiency and reliance on sunlight and the

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wind. As flow resources, solar and wind energy are available in large, undiminishable quantities. But exploiting these resources in a cost-effective way depends critically on both the state of technology and investments in the needed infrastructure and equipment. In this analysis, the sustained availability of fossil fuels would be unnecessary to equip future generations to thrive and flourish. One final topic concerns the interplay between questions of intra- and intergenerational fairness. On the one hand, utilitarians would attach equal weight to the interests of each present and future person, suggesting a need to level inequalities within generations while stepping up rates of capital investment to maximize total welfare in the long run. Whether or not this is plausible is an open question. In rights-based ethics, on the other hand, concepts of intergenerational justice can be derived from the notion that each member of society should have an equal opportunity to define and pursue their interests (Howarth, 1992). This is generally consistent with the Capabilities Approach described above and the arguments set forth in Sen (2009). In practice, however, actions that benefit future generations can impose hardships that exacerbate the challenges faced by people who are socially and economically disadvantaged. This challenge can arise with regard to programs aimed at conserving ecosystems and increasing carbon storage in the Global South, as well as actions that negatively impact less privileged groups in the Global North. Such conflicts can sometimes be resolved by integrating the insights produced by the fields of ecological economics and political ecology (Martinez-Alier, 2002; Sneddon et al., 2006). Yet the challenges are not trivial, noting that they often intersect with questions concerning the autonomy and fair treatment of Indigenous people, women, and ethnic minorities. Richard B. Howarth

References

Boulding, K.E. 1966. “The Economics of the Coming Spaceship Earth.” In H. Jarrett (ed.), Environmental Quality in a Growing Economy, 3–14. Baltimore, MD: Johns Hopkins University Press. Broome, J. 2008. “The Ethics of Climate Change.” Scientific American 298: 97–102.

Cline, W. 1992. The Economics of Global Warming. Washington, D.C.: Institute for International Economics. Daly, H.E. 1977. Steady-State Economics. San Francisco: W.H. Freeman. Dasgupta, P.S., and G.M. Heal. 1979. Economic Theory and Exhaustible Resources. Cambridge: Cambridge University Press. Ekins, P., S. Simon, L. Deutsch, C. Folke, and R. De Groot. 2003. “A Framework for the Practical Application of the Concepts of Critical Natural Capital and Strong Sustainability.” Ecological Economics 44: 165–85. Howarth, R.B. 1992. “Intergenerational Justice and the Chain of Obligation.” Environmental Values 1: 133–140. Howarth, R.B. 2011. “Intergenerational Justice.” In J.S. Dryzek, R.B. Norgaard, and D. Schlosberg (eds.), Oxford Handbook on Climate Change and Society, 338–53. Oxford: Oxford University Press. Koopmans, T.C. 1972. “Representation of Preference Orderings Over Time.” In C.B. McGuire and R. Radner (eds.), Decision and Organization: A Volume in Honor of Jacob Marschak, 103–24. Amsterdam: North Holland. Martinez-Alier, J. 2002. The Environmentalism of the Poor: A Study of Ecological Conflicts and Valuation. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. https://​www​.e​ -elgar​.com/​shop/​gbp/​the​-environmentalism​-of​ -the​-poor​-9781840649093​.html Nordhaus, W.D. 2007. “A Review of the Stern Review on the Economics of Climate Change.” Journal of Economic Literature 45: 686–702. Norton, B.G., and M.A. Toman. 1997. “Sustainability: Ecological and Economic Perspectives.” Land Economics 73: 553–68. Parfit, D. 1984. Reasons and Persons. New York: Oxford University Press. Ramsey, F.P. 1928. “A Mathematical Theory of Saving.” The Economic Journal 38: 543–59. Raworth, K. 2017. Doughnut Economics. White River Junction, VT: Chelsea Green Publishing. Rockström, J., W. Steffen, K. Noone, Å. Persson, F.S. Chapin, III, E. Lambin, T.M. Lenton, et al. 2009. “Planetary Boundaries: Exploring the Safe Operating Space for Humanity.” Ecology & Society 14: 32. Sen, A. 2009. The Idea of Justice. Cambridge, MA: Harvard University Press. Sidgwick, H. 1901. The Method of Ethics. London: MacMillan. Singer, P. 2002. One World: The Ethics of Globalization. New Haven, CT: Yale University Press. Sneddon, C., R.B. Howarth, and R.B. Norgaard. 2006. “Sustainable Development in a Post-Brundtland World.” Ecological Economics 57: 253–68.

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University Press. World Commission on Environment and Development. 1987. Our Common Future. London: Oxford University Press.

48. Georgescu-Roegen’s bioeconomics 48.1 Introduction

A divergent series of his work, Georgescu-Roegen’s epistemological reflections cover various economic subjects. These preoccupations served as the main source of inspiration for his entire corpus of scientific contributions. They particularly inspired Georgescu-Roegen toward his firm belief that any theory must be an operational description of a reality’s mode of functioning.

48.2 Georgescu-Roegen’s epistemology

Georgescu-Roegen’s (1971) epistemology is concerned with three inquiries: (1) what the essence of theoretical science is; (2) whether or not economics is a theoretical science; and (3) what the proper approach to economic science is. Theoretical science concerns logically ordered descriptions derived from noncontradictory axioms. Newtonian mechanics is an example of such a science that, after having reached a supreme position in physics, was abandoned in the wake of three major discoveries: the entropy law in classical thermodynamics, quantum theory, and relativity. In theoretical science, such revolutions are always brought about by an introduction of new axioms from which proper deductions are made so as not to cause any contradiction to confirmed empirical findings. The principle of contradiction—B cannot be both B and non-B—is the essence of theoretical science. The principle of contradiction prevails in the real number system (i.e., each number is discretely distinct). Georgescu-Roegen referred to models using discretely distinct concepts conforming to the principle of contradiction as arithmomorphic. So, in any theoretical science, a sort of tautology prevails until new empirical evidence, not consistent with the preexisting axioms, is ascertained. The role of arithmomorphic models is to facilitate argumentation, clarify results, and so guard against possible faults of analytical reasoning. A representative example of arithmomorphic models is a dynamical model of mechan-

ics, extensively used in standard economics. Standard economics still clings to the mechanistic epistemology, in which an objective function is optimized subject to a constraint. In mechanical systems, energy and matter entering a process must come out in exactly the same quantity, following the principle of conservation. Naturally, the analysis of mechanics, concerned with changes in locomotion, often leads to unrealistic descriptions and conclusions when applied to the economic process: (1) the economic process is reversible; (2) the future path of the economic process is predictable; (3) variables in the economic process are under human control; (4) an initial equilibrium is always recoverable after a perturbing external force is removed; and (5) the economic process is circular and cannot possibly affect the environment. Georgescu-Roegen’s corpus stands in opposition to the mechanistic epistemology, resorting to the entropy law as a theoretical edifice. Life is, in its essence, an entropic process. Organisms maintain their own highly ordered structures by sucking low entropy from the environment to compensate for continuous internal entropic degradation. From a biophysical perspective, the economic process is also an irreversible entropic process, taking in low-entropy energy and matter and disposing high entropy energy and matter into the environment. Taking to heart these observations, a completely different view on the economic process is in order: (1) the economic process is irreversible; (2) the future path of the economic process is uncertain, due to entropic indeterminacy (Georgescu-Roegen 1971); (3) the economic process is not always under human control; (4) the economic process is constantly changing and often punctuated by unexpected events; and (5) the economic process is not an isolated, self-sustaining process and cannot go on without a continuous energy and material exchange with the environment, irrevocably altering the environment. Recognizing these points, economists must be prepared to squarely face inevitable qualitative changes in the economic process. There commonly arise situations in the economic process where the principle of contradiction cannot be applied—situations where both A and non-A overlap. Georgescu-Roegen proposed to call the concepts that violate the principle of contradiction dialectical, where

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dialectical concepts can be seen to reflect the most essential aspects of change. Correct reasoning with proper use of dialectical concepts is indispensable for economic science. Though such reasoning is a far more delicate operation than its arithmomorphic counterpart, dialectical reasoning allows the user to avoid formalism nonsense and abuse of empty mathematical exercises.

48.3

Analysis combined with dialectics

Ecological economists do not seem to sufficiently recognize Georgescu-Roegen’s analytical and dialectical precision of thought, which can be seen to surpass a vast majority of the economic writings. It is worthwhile to present Georgescu-Roegen’s related deepest results, which exemplify the precision of his trailblazing contributions. 48.3.1 Cardinal measurability One belief commonly held in economics is that ordinal measurability, that which is indifferent to monotonic transformation of scale, is sufficient for economic science. If we accept that any given scale is as good as any other, such fundamental notions of decreasing marginal scale, for example, lose all meaning. In this way, ordinal measurability can be seen as entirely insufficient when used to describe the production process. Cardinal measurability (e.g., weight) is the result of a series of specific physical operations without which mathematical manipulations would have no relevance. It reflects a particular physical property of a category of things. Georgescu-Roegen (1964) established five axioms, enumerated A1, A2, A3, A4, and A5 below, which are a minimal set of independent axioms required to establish a cardinal measure: ● A1. The set C is subsumptive: C has a subsumption operation that is not dependent of the order of operations, and C has the neutral element, which is “nothing.” Colors can be subsumed by mixing, but the set of all colors has no neutral element. If an element a is a subsumption of a pair of two elements b and c, and c is not the neutral element, a is defined to be greater than b (i.e., a G b). A1 does not exclude the possibility that there is a pair Kozo Torasan Mayumi

of elements for which both a G b and b G a are satisfied. ● A2. The set C has a property that, if a subsumption of two elements is equal to one of the two elements, then one of the two elements is the neutral element. A2 signifies the absoluteness of the notion of “nothing.” ● A3. If a subsumption of two elements is the neutral element, both of the two elements must be the neutral element. A3 means that the neutral element cannot be greater than any other element. It is easy to show that A3 cannot be derived from A1 and A2, and that A2 cannot be derived from A1 and A3. Thus, A1, A2, and A3 are not redundant. Yet, A1, A2, and A3 do not guarantee to render any two elements comparable in terms of the relation G. We require another axiom that guarantees that all of its elements are comparable through the relation G. Such an axiom is the homogeneity of elements, expressed by A4: ● A4. Given any pair of elements of a set, at least one element can be found in the set for which one element of the pair is greater than the other element of the pair. Note that A1 and A4 can guarantee neither A2 nor A3. A4, the homogeneity axiom, allows for the operation of subtraction. Then, a new relation Γ can be obtained from the relation G: a Γ b means either a = b or a G b. A set satisfying A1, A2, A3, and A4 can be ordered by the relation Γ. On the other hand, not every set satisfying A1, A2, A3, and A4 is ordinally measurable. Lexicographic order is not ordinally measurable, so it is impossible to have a linear continuum—that is a basis of arithmomorphic models. The Lexicographic order example shows the ordinalist fallacy: comparability does not always lead to ordinal measure. In fact, there is a missing axiom, the Archimedean property, which produces a cardinal measure. This property is included in the completeness of the real numbers: ● A5. The set C has the Archimedean property when, given two elements a and b in C, a not being the neutral element, then, there exists a natural number k such that ka G b.

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In plain terms, we must empty a reservoir by removing water over finite steps. It must now be noted, though without proof here, that A1, A2, A3, and A5 alone cannot produce a set having an ordinal measure. Thus, both A4 (the homogeneity axiom) and A5 (the Archimedean property) are indispensable for a set to be a cardinal measure. The present author believes that, after an investigation on measurability in economics, Georgescu-Roegen encountered Lexicographic ordering where arithmomorphic representation is impossible because of qualitative variations in the object of preference. Georgescu-Roegen reconsidered the nature of economic science that treats a variety of qualities that cannot be represented by a linear continuum, so that he tried to look for an alternative approach to economics beyond the method based on arithmomorphic representation, typical of Newtonian mechanics. 48.3.2 Defects of the Leontief dynamic model The Leontief dynamic model is frequently used by planning agencies for economic development. As Georgescu-Rogen (1976, Chapter 9) noted, however, that model rests on unrealistic assumptions. There are three major defects associated with the Leontief dynamic model. First, it has a serious inconsistency: at the initial point of time, it is required that every recipe in a particular process requires a certain amount of goods produced by one or more of the other processes. This conundrum, how the initial production process has come to be in the first place, cannot be solved within the Leontief model itself. Creation of processes is a prerequisite for commodity production. The Leontief model cannot throw any light on economic planning from the get-go. This first problem is related to the second problem: the model’s inherent quasi-explosive feature. The quasi-explosive feature refers to the situation where, as soon as the necessary funds have been procured to expand production, the net product level instantaneously jumps to a higher level. Funds are agents transforming a given set of inflows into a given set of outflows. They are the elements that enter and leave the process unchanged during a transformation process: labor, capital, and Ricardian land (Georgescu-Roegen, 1971).

The net product starts to increase at the very moment the previous consumption level is decreased. In plain terms, a miracle occurs! Unless the capacity of funds is sufficient for increasing production, some additional processes must first be created to increase product flow. Any process can start producing a product outflow only after the process is primed (i.e., only after all its process-funds are completed). Both to build a process out of commodities and to prime the process requires some duration in addition to the time necessary to accumulate funds. Time lags are absolutely necessary to expand production (see Mayumi, 2001, Chapter 3). Third, decumulation of capital is not accumulation of capital in reverse order. That is termed relaxation phenomena, asymmetric periodicity (Georgescu-Roegen, 1966, Chapter 8). The Leontief model cannot deal with two different phases within the same representation. 48.3.3 Critique of Marx and Arrow-Debreu on agrarian problems Georgescu-Roegen (1960) confirmed that both Marxist and Standard Theory is far from satisfactory for the analysis of an overpopulated agricultural economy. As Mitrany (1951) observes, Marx’s view on peasant agriculture combines “the townman’s contempt for all things rural and the economist’s disapproval of small-scale production” (6). Marx assumed that “the loss to the peasant would be the gain of society” (13). On the other hand, standard analysis of marginal productivity does not much help, either. Georgescu-Roegen showed that agrarian economies cannot possibly function according to the principle of marginal productivity. He correctly stated that agrarians have never lost sight of the most elementary principle of economic development: no factor should remain unnecessarily idle. In overpopulated agrarian economies, this means using labor even to the point where its marginal productivity becomes zero, resulting in violating the marginal principle adopted in standard economics. He also argued that most underdeveloped agrarian economies are poor due to relative land shortage as well as to a chronic dearth of capital. At this juncture, in relation to agrarian issues, Georgescu-Roegen criticizes Arrow-Debreu’s celebrated paper. The economic problem is set aside by assumKozo Torasan Mayumi

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ing that every member of the community is endowed ab initio with a real income sufficient for their entire life span. In actual life, people are endowed with labor of limited capacity and use resources of limited quantities. Under such conditions, the solution by Arrow and Debreu cannot be reached by the mechanism of marginal principles.

Georgescu-Roegen’s condition is weaker and more economical in terms of computational time complexity: if the number of equations is n, then the Hawkins–Simon conditions require a time order of 2n while the Georgescu-Roegen conditions require a time order of only n to calculate all required determinants.

48.3.4 Issues of dimension Georgescu-Roegen published two papers concerned with dimensional issues (Pigou et al., 1936; Georgescu-Roegen, 1966, Chapter 12). The first paper is a verdict on a controversy between Pigou and Friedman concerning the elasticity of demand. Friedman adopted a strict interpretation of mathematical constant, while Pigou considered a nearly constant elasticity of demand. Pigou’s method is theoretically and practically correct as far as a small part of income is spent on any commodity. Furthermore, what Georgescu-Roegen proved is the absurdity of qualifying a dimensional entity as small or as large. The verdict was against Friedman. In fact, Georgescu-Roegen showed how the assumption of a very large number of commodities may lead to either a zero or a finite value for the elasticity of utility of money with respect to the price of a commodity, depending on the type of utility function adopted (Pigou et al., 1936). The second paper shows the logical fallacy of the Marxian formulation of expanded reproduction. Such a formulation violates the principle of dimensional homogeneity: for example, l and dl/dt cannot be added where l represents the consumption of capitalists’ households, if t has dimension of time adopted in the Marxian formulation (Georgescu-Roegen, 1966, Chapter 12). Georgescu-Roegen successfully constructed a dynamic scheme free from dimensional contradictions but embodied as many essential points of the Marxist rationale as possible, and then discussed the central theme of the inadequacy of the capital accumulation process in the capitalist system.

48.4

48.3.5 Hawkins–Simon condition Georgescu-Roegen (1950) investigated similar conditions to those of Hawkins and Simon (1949). Hawkins–Simon conditions are concerned with the necessary and sufficient conditions for the existence of static equilibrium for a Leontief limitational model. Kozo Torasan Mayumi

The culmination to Georgescu-Roegen’s bioeconomics

Human species transgressed the mode of endosomatic evolution by entering into a far faster evolutionary mode where exosomatic organs are manufactured. The institutions of the market, money, credit, and enterprises of all sorts emerged in response to this progressive evolution. It is true that economic growth and the advancement of science and technology through exosomatic evolution resulted in the apparent increased material comfort attained by the Western world. Yet, exosomatic evolution brought to the forefront of humanity three formidable predicaments to human beings. The first predicament concerns the eventual exhaustion of fossil fuels and mineral resources associated with humanity’s addiction to exosomatic instruments. Steam engines, and later but much more importantly, internal combustion engines, are prime technological wheels of the modern civilization. These two types of technologies have driven a common explosive characteristic typical of “Promethean technologies”: more and more coal and oil extraction, more than what was previously used in the entirety of the economic process. Due to its explosive nature and the related depletion of fossil fuels, Promethean technology has driven humans at full speed into the Malthusian instability trap. In particular, the explosive characteristic of the petroleum-based metabolism of modern society, due to the abundant supply of high-quality oil in the past hundred years or so and the continuous technological efficiency improvements, have been boosting the phenomena associated with Jevons’ paradox. Neither solar energy nor thermonuclear energy would be able to support, easily and on a grand scale, current fossil-fuel-based economies. Fossil fuels are optimal support-

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ers of modern industrial society in terms of the amount of matter in bulk required for energy extraction, transformation, and transportation. This proposition can be termed Georgescu-Roegen’s Fundamental Proposition. In the case of solar energy at the ground level, the necessary amount of matter is high for weak-intensity energy because such energy must be concentrated into a much higher intensity if it is to support intensive industrial processes such as those supported by fossil fuels. In the case of thermonuclear energy, the necessary amount of matter is also high for high-intensity energy because high-intensity energy must be contained and controlled within a stable boundary. The second predicament is endless social conflict. Indeed, there are other species living in societies, but without such conflict. Their social classes correspond exactly to somatically inherited roles without causing any conflict. Phragmotic-headed ants, for instance, prevent intruders from entering nests by blocking the entrances with specially modified body structures. In contrast to the case of phragmotic-headed ants, social conflict over exosomatic instruments does characterize human species. As Lotka (1956) clearly recognized, the fact that natural resources—in particular, accessible energy and minerals—are unevenly distributed has led to “so much of the social unrest that has accompanied the development of modern industrialism” (370). Since large-scale production and its distribution must be organized socially, the social classes of “ruler” and “ruled” are created. Unfortunately, social conflict will remain part of the human lot as long as existence depends on large-scale exosomatic production and distribution. The root point is that no social system, socialism or what have you, can bring about a New Jerusalem. Georgescu-Roegen’s view of social conflict is a strong rebuttal to Marx’s thesis of historical materialism, Marxian mechanistic determinism. The last predicament has to do with inequality among different exosomatic “species,” for example, the difference between the developed and underdeveloped countries. This, frankly, very sad predicament is an intragenerational issue, that is becoming a crucial event in economically advanced societies due to the extraordinary expansion of financial markets where earning unearned income is

a sort of a golden calf. Concern for sustainability invokes another type of distributional issue—the intergenerational distribution issue. The notion of intergenerational distribution is strongly related to the selection of a proper discount rate, if any, and to the notion of sustainability, weak or strong. Since humans seem to be so enchanted with miraculous technological achievements, there is difficulty in recognizing that exosomatic evolution has unavoidably brought about these three lasting predicaments. The scope of economics is not confined to the study of how given means are applied to satisfy given ends, reducing economics to the mechanics of utility and self-interest. New means are continually invented, new economic wants created, and new distributive rules introduced—those are the essence of economic change. The essence of economic change consists of the organizational and flexible power to create new processes rather than the power to produce commodities. Georgescu-Roegen (1971) calls such power a Π-sector. An economy can “take off” when, and only when, it has succeeded in creating a Π-sector. This issue of the Π-sector is related to the issue of what is produced in the economic process. Some who study the functioning of socioeconomic processes seem to be confused by what is produced by the economic process. Commodities are not produced by commodities except in a stationary state. The economic process has the goal of reproducing and expanding the various fund elements defined simultaneously across different levels and scales, by using disposable flows. Following Georgescu-Roegen, we can conclude that an economy not only produces goods and services but, more importantly, produces the processes required for producing and consuming goods and services. Rapid expansion of the Π-sector, a fountainhead of economic development, dictates an increase not only in energy requirement, but also in mineral resources demand. Flows of dissipated matter in bulk increase with the size of the economic process, and there is great difficulty in maintaining these large-scale material structures of modern industrial society. Georgescu-Roegen’s concern with mineral resources led him to propose his fourth law of thermodynamics: matter also irrevocably decays (Georgescu-Roegen, 1977). It is none other than Clausius (1862) Kozo Torasan Mayumi

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who proposed to measure the degree of dispersion of matter in terms of a then new variable: disgregation. Clausius showed that entropy is the sum of disgregation and heat dispersion. Taken independently, neither component of entropy is a total differential. In other words, neither the integral of disgregation nor the integral of heat dispersion can be transformed into a thermodynamic state function. This is a critical insight that explains why Georgescu-Rogen’s fourth law cannot be established as a physical law. The fourth law refers only to the disgregation component of entropy, while modern physicists generally refuse to see that entropy is concerned with both energy and matter dispersion. To such physicists, it should be noted that Planck (1945) himself states: The real meaning of the second law has frequently been looked for in a “dissipation of energy.” This view, proceeding as it does from the irreversible phenomena of conduction and radiation of heat, presents only one side of the question. There are irreversible processes in which the final and initial states show exactly the same form of energy (e.g., the diffusion of two perfect gases) or further dilution of a dilute solution. Such processes are accompanied by no perceptible transference of heat, nor by external work, nor by any noticeable transformation of energy. They occur only because they lead to an appreciable increase of the entropy. In this case it would be more to the point to speak of a dissipation of matter than of a dissipation of energy. (104)

So, Planck vindicates that neither physicists nor Georgescu-Roegen are correct. Both sides only tell one part of the true meaning of entropy.

48.5 Conclusion

This entry presents a summary of Georgescu-Rogen’s works, all-emergent from his sincere epistemological preoccupation in science, culminating in his own bioeconomics proposal. As far as the author is aware, Georgescu-Roegen is the only econ-

Kozo Torasan Mayumi

omist who states, without hesitation, that the primary purpose of economic activity is the self-preservation of the human species. Self-preservation requires the satisfaction of basic needs. Of course, in maintaining all the various necessities of a decent life, the purely biological ones are absolutely indispensable for survival. Since biological life feeds on low entropy, Georgescu-Roegen identified and emphasized the most important connection between economics and low-entropy resources. Georgescu-Roegen’s penetrating analytical ability also allowed him to identify the analytical fallacy of the standard economists’ mechanistic epistemology. Hence, it is no wonder that Georgescu-Roegen was never awarded the Sveriges Riksbank Prize in Economic Sciences. For standard economists, to give such an award to Georgescu-Roegen would stand as a self-signing of their own death sentences, clearly impossible for them to accept! On the other hand, as the last student of Georgescu-Roegen, I believe he must have been very proud of having never been awarded such a controversial prize. I do hope that the future generation of economists will learn to speak, fluently, Georgescu-Roegen’s language, toward achieving a society aiming at self-preservation of the human species. Kozo Torasan Mayumi

References

Clausius, R. 1862. “On the application of the theorem of the equivalence of transformations to interior work.” In T. A. Hirst (ed.), The Mechanical Theory of Heat with Its Applications to the Steam-Engine and to the Physical Properties of Bodies. London: John van Voorst, 215–50. Georgescu-Roegen, N. 1950. “Leontief’s system in the light of recent results.” The Review of Economics and Statistics 32: 214–22. Georgescu-Roegen, N. 1960. “Economic theory and agrarian economics.” Oxford Economic Papers 12: 1–40. Georgescu-Roegen, N. 1964. “Measure, quality, and optimal scale.” In C. R. Rao (ed.), Essays on Econometrics and Planning Presented

Georgescu-Roegen’s bioeconomics  283 to Professor P. C. Mahalanobis. Oxford: Pergamon, 231–56. Georgescu-Roegen, N. 1966. Analytical Economics, Cambridge, MA: Harvard University Press. Georgescu-Roegen, N. 1971. The Entropy Law and the Economic Process, Cambridge, MA: Harvard University Press. Georgescu-Roegen, N. 1976. Energy and Economic Myths, New York: Pergamon Press. Georgescu-Roegen, N. 1977. “The steady state and ecological salvation: A thermodynamic analysis.” BioScience 27: 266–70. Hawkins, D., and Simon, H. A. 1949. “Note: Some conditions of macroeconomic stability.”

Econometrica 17: 245–8. Lotka, A. J. 1956. Elements of Mathematical Biology. New York: Dover Publications. Mayumi, K. 2001. The Origins of Ecological Economics: The bioeconomics of Georgescu-Roegen. London: Routledge. Mitrany, D. 1951. Marx Against the Peasant: A Study in Social Dogmatism. Chapel Hill: University of North Carolina Press. Pigou, A. C., Friedman, M., and Georgescu-Roegen, N. 1936. “Marginal utility of money and elasticities of demand.” Quarterly Journal of Economics 50: 532–9. Planck, M. 1945. Treatise on Thermodynamics, 7th ed. New York: Dover.

Kozo Torasan Mayumi

49. Green economy Ecological economics views the economy as a subsystem of the planetary ecological system, fundamentally limited by the physical realities of that planetary system. This has always been true, but it was possible to neglect the implications of this basic truth so long as human economic activity was at a relatively low level relative to planetary capacity—allowing economic theorists to take what Herman Daly (1974, 1993) has referred to as an “empty world” rather than a “full world” perspective. This perspective, typical of neoclassical economics, essentially ignores or minimizes resource and environmental constraints. In the 21st century, this approach is no longer possible. Since about 1950, there have been substantial increases in the global population, energy use, and carbon emissions—more than threefold for the global population and more than sixfold for energy use and carbon emissions. This has led to an intensifying climate crisis, and parallel crises have emerged in other resource and environmental areas, including issues of overdraft of water resources, decline of forests and wetlands, degeneration of agricultural soils, ocean pollution, fisheries decline, and biodiversity loss (United Nations Environment Programme [UNEP], 2019; Intergovernmental Panel on Climate Change, 2021). Even with optimistic forecasts of population stabilization, these consumption-generated pressures on the global ecosystem can be reliably forecast to increase further during the 21st century. An ecological economics perspective thus implies a drastic change in the nature of economic production, as the period of steady economic growth—characteristic of the past 200 years and especially of the last 75 years— necessarily comes up against firm ecological limits. Changes in production systems that are essential for a “green” economy include: ● Shifting agricultural practices away from current patterns of heavy reliance on chemical inputs; development of regenerative agricultural systems to build up soil productivity and carbon content, conserve water, and integrate crop production with agroforestry and sustainable livestock systems.

● Eliminating fossil fuel dependence through a rapid transition to renewable energy sources and energy efficiency, with a goal of net-zero carbon emissions. ● Development of “circular economy” practices for reuse and recycling of resources, with elimination of chemical and plastic waste entering the environment. ● Sustainable management of forests and wetlands, preserving or expanding current area in forests and wetlands while eliminating destructive harvest practices, and significantly expanding forest carbon storage. ● Sustainable management of fisheries, including consideration of impacts on non-harvested species. ● Conservation of species diversity, with major areas being set aside for conservation and modified management of “buffer” areas to protect ecosystems and species diversity. An important issue is whether these changes are possible in the context of continuing economic growth. In theory, “absolute decoupling”—reducing overall resource inputs, specifically carbon-based fuels—may be possible while still “growing” the economy. Advocates of “degrowth” argue that absolute decoupling is unlikely, meaning that consumption in developed nations must be reduced if carbon reduction targets are to be achieved. “Relative decoupling”—reducing the carbon intensity of the economy—has been adopted as a goal by some major developing economies, such as China and India, where consumption is still growing. Both in developed and developing economies, major changes will be required to redirect economic activity away from a carbon-intensive path. A similar logic applies to other major ecological constraints, including water supply, soil fertility, species diversity, and ocean ecosystems (Harris, 2019a). The UNEP (2011, 2018) green economy initiative, “promoting the transition to economies that are low-carbon, resource-efficient, and socially inclusive” includes policy recommendations, such as: ● Using taxes and other market-based instruments to internalize negative externalities. ● Decreasing government spending and subsidies that deplete natural capital.

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● Implementing rigorous efficiency and technology standards. ● Phasing out polluting industries while developing and phasing in green technologies. ● Supporting an employment transition for affected workers, providing training for displaced workers to gain new jobs in the green economy. ● Strong international agreements to deal with global environmental issues, such as climate change and ozone depletion. Green economy initiatives have been taken up in the United States and the European Union, under the rubric of the Green New Deal (GND) and the European Green Deal. The GND takes its inspiration from the original New Deal under Franklin Roosevelt, which countered mass unemployment in the 1930s. It envisions an ambitious national mobilization to achieve net-zero greenhouse gas emissions through investments in energy efficiency, renewable energy, zero-emission vehicles, high-speed rail, and other infrastructure, with a goal of achieving net-zero emissions no later than 2050. Economists who have analyzed the Green New Deal note that cost and benefit estimates depend on which specific proposals are included. Initial costs for many aspects of the program may be quite low, since energy efficiency and renewable energy are already very economically viable (Harris, 2019b). Getting to the goal of net-zero emissions is more difficult. Robert Pollin (2015) proposes that reaching 80 percent renewables by 2035 and 100 percent renewables by 2050 is “realistic if very, very challenging,” and would be compatible with the Paris Climate Agreement targets. Edward Barbier calls for the U.S. to invest about $200 billion annually to address climate change, funded by a carbon tax and elimination of fossil fuel subsidies, and possibly a tax on the highest-income earners (Barbier, 2019). Renewable energy is not a panacea, since “renewable” refers only to the energy source and not to the materials that are required to build the structures (e.g., solar arrays and wind turbines). These materials are not renewable and only partially recoverable, reusable, and recyclable. Thus, a transition to renewables needs to be accompanied by the expansion

of methods for net carbon removal, whether nature-based or artificial. The greatest known potential lies in carbon storage in agricultural soils, forests, and wetlands; artificial carbon removal is more speculative. A key element of GND plans is job creation. A transition away from fossil fuels means that jobs will be lost in the coal, oil, and gas sectors, but new job creation in other sectors, such as solar energy, is expected to be much larger. This still leaves a significant problem of transitional aid, since those who lose jobs in one sector will not necessarily be able to regain them in another. In some cases retraining is required; in other cases different kinds of job creation, for example, in environmental restoration of areas damaged by strip mining, or expansion of rural health care, may be appropriate (Brown and Ahmadi, 2019). The European Green Deal has a goal of net-zero greenhouse gas emissions by 2050. Strategies include decarbonizing the energy sector, expanding public transit, promoting energy efficient buildings, and investing in environmentally friendly technologies. The EU also promises to “provide financial support and technical assistance to help those that are most affected by the move towards the green economy” (European Commission, n.d.). The concept of a “green economy” implies that improved human well-being and reduced inequality can be achieved through investments that reduce environmental impacts and promote a transition to sustainable forms of production. The common perception that protecting the environment is detrimental to the economy is not borne out by numerous studies. Strong environmental regulations often involve some economic costs, but evidence indicates that the benefits of environmental regulations far exceed their costs. Rather than leading to job losses, protecting the environment through well-designed policies can actually be a source of net job creation (Harris and Roach, 2022). While creating a green economy will entail short-term costs, the long-term economic and environmental benefits are projected to be much greater. The transition to a green economy will require strong policy action, including increasing investment in renewable energy and regenerative agriculture, eliminating harmful subsidies, retraining Jonathan M. Harris

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workers for green industries, using economic policy instruments such as taxes and tradable permits, and strengthening international agreements that protect the environment (Richardson, 2013). Whether a green economy is compatible with “green growth” is controversial. Some analysts believe that the green growth concept makes sense, provided that natural capital—the resources and environmental services that support the economy—can be fully accounted for and maintained: “Moving to inclusive green growth necessitates a rigorous understanding of the significance and values of natural capital for human well-being. Whether in the public or private sectors, decision-makers will need to be persuaded of the benefits from major investments in nature and nature-based solutions compared to their costs” (Mandle et al., 2019, 4). Others argue that “green growth” is an oxymoron: “There’s no question that economic growth depends crucially on what nature provides for human economies to function and grow . . . there is compelling evidence that economic growth is stressing nature’s capacity beyond its limits” (Victor, 2019, 94).

Source:

Harris and Roach (2022, Chapter 22).

Figure 49.1 An economy experiencing degrowth

Jonathan M. Harris

One review of the literature on the relationship between economic growth and sustainability concludes that “green growth paths are unlikely to be sustainable,” mainly because carbon emissions will fail to fall quickly enough to meet global climate objectives (Kallis et al., 2018, 308). A review of decoupling studies reached a similar conclusion, detecting no “economy-wide resource decoupling, neither on national nor international scales.” The authors recommend that “more attention should be given to conceptualisations of economy that do not rely on economic growth as the key route towards ecological sustainability and human wellbeing” (Vadén et al., 2020, 243). Given the limitations of green growth, an alternative path to sustainability is a fundamental reorientation of developed economies away from continual economic and material growth. This path was first advocated by Herman Daly, one of the founders of ecological economics, in the 1970s, with his proposal for a steady-state economy (SSE). Daly defined the three main characteristics of an SSE as: ● a stable human population; ● a stable stock of physical resources; and

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● a minimum throughput (combined input and output) of materials and energy. Other ecological economists have called for a transition to a “post-growth” economy that would emphasize sustainability rather than further economic growth (Jackson, 2017). An economy experiencing degrowth is presented in Figure 49.1. As this hypothetical economy experiences exponential economic growth, it eventually outstrips its ecological carrying capacity, causing a reduction in ecological resilience. The size of the economy reaches a peak, and then a period of degrowth occurs, either because of conscious policy, or due to an “overshoot and collapse” syndrome. Eventually the ecosystem stabilizes at a new, though lower, carrying capacity. Once the size of the economy again falls within carrying capacity, an economic steady state can be maintained. A particular steady-state level need not be indefinite—as ecological conditions change, the level of economic activity can be adjusted upward or downward as needed. A possible resolution of the “green growth” versus “degrowth” argument could be to apply the degrowth principle only to certain resource and energy-intensive sectors of the economy. Reducing or eliminating consumption of fossil fuel energy, resource-intensive agriculture, single-use plastics, and so on, could allow for expanding regenerative agriculture, renewable energy, social services, health, education, and other areas. In this way, the effect on overall economic growth would be indeterminate, depending on the relative rates of decrease and increase in output of different economic sectors. In general, a focus on sustainable systems and human services would be associated with more labor-intensive methods, leading to higher employment. This could reduce gross domestic product (GDP) growth as traditionally measured, but numerous alternative measures of economic welfare indicate that GDP is not the optimal measure of economic success (Harris and Roach, 2022). A green economy would therefore be consistent with expanded employment and improved human well-being. Jonathan M. Harris

References

Barbier, Edward. 2019. America Can Afford a Green New Deal – Here’s How. The Conservation, February 26. Brown, Marilyn, and Majid Ahmadi. 2019. Would a Green New Deal Add or Kill Jobs? Scientific American, December 17. Daly, Herman. 1974. The Economics of the Steady State. American Economic Review, 64(2):15–21. Daly, Herman. 1993. The Steady-State Economy: Toward a Political Economy of Biophysical Equilibrium and Moral Growth. In Herman Daly and Kenneth Townsend, eds., Valuing the Earth, 2nd ed. Cambridge, MA: MIT Press, 325–64. European Commission. n.d. A European Green Deal: Striving to be the First Climate-Neutral Continent. https://​ec​.europa​.eu/​info/​strategy/​ priorities​-2019​-2024/​european​-green​-deal​_en Harris, Jonathan M. 2019a. Responding to Economic and Ecological Deficits. Tufts University Global Development and Environment Institute Working Paper No. 19-01, April. https://​sites​.tufts​.edu/​gdae/​ working​-papers Harris, Jonathan M. 2019b. Ecological Economics of the Green New Deal. Tufts University Climate Policy Brief #11, August. https://​sites​ .tufts​.edu/​gdae/​climate​-policy​-briefs/​ Harris, Jonathan M., and Brian Roach. 2022. Environmental and Natural Resource Economics: A Contemporary Approach, 5th ed. New York and London: Routledge Intergovernmental Panel on Climate Change. 2021. Climate Change 2021: The Physical Science Basis. https://​www​.ipcc​.ch/​report/​ar6/​ wg1/​ Jackson, Tim. 2017. Prosperity Without Growth: Foundations for the Economy of Tomorrow. New York and London: Routledge. Kallis, Giorgos, Vasilis Kostakis, Steffen Lange, Barbara Muraca, Susan Paulson, and Matthias Schmelzer. 2018. Research on Degrowth. Annual Review of Environment and Resources, 43:291–316. Mandle, Lisa, Zhiyun Ouyang, James Salman, and Gretchen C. Daily, eds. 2019. Green Growth that Works: Natural Capital Policy and Finance Mechanisms around the World. Washington, D.C.: Island Press. Pollin, Robert. 2015. Greening the Global Economy. Cambridge, MA: MIT Press. Richardson, Robert B., ed. 2013. Building a Green Economy: Perspectives from Ecological Economics. East Lansing: Michigan State University Press. United Nations Environment Programme (UNEP). 2011. Towards a Green Economy: Pathways

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288  Elgar encyclopedia of ecological economics to Sustainable Development and Poverty T. Toivanen, E. Hakala, and J.T. Eronen. 2020. Eradication. Nairobi: UNEP. Decoupling for Ecological Sustainability: UNEP. 2018. Green Industrial Policy: Concepts, A Categorisation and Review of Research Policies, Country Experiences. http://​ www​ Literature. Environmental Science and Policy, .unep​.org/​explore​-topics/​green​-economy 112:236–44. UNEP. 2019. Global Environment Outlook Victor, Peter A. 2019. Managing Without – GEO-6: Healthy Planet, Healthy People. Growth: Slower by Design, not Disaster, 2nd Nairobi: UNEP. ed. Cheltenham, UK, and Northampton, MA: Vadén, T., V. Lähde, A. Majava, P. Järvensivu, Edward Elgar Publishing.

Jonathan M. Harris

50. Human appropriation of net primary production (HANPP) 50.1 Introduction

50.2

In many regions, humanity’s impact on the biosphere’s structures (e.g., land cover) and functioning (e.g., biogeochemical cycles) already exceeds natural variability (Steffen et al., 2007). Up to 83 percent of the global terrestrial biosphere has been classified as being under direct human influence, based on geographic proxies such as human population density, settlements, roads, or agriculture (Venter et al., 2016). Humans use approximately three-fourths of the entire global land mass, except Greenland and Antarctica (Arneth et al., 2020; Ellis et al., 2013). The human appropriation of net primary production (HANPP) is an aggregated indicator that reflects both the area of land used by humans and the intensity of the respective land uses. HANPP measures to what extent land conversion and use as well as biomass harvest alter the availability of trophic (biomass) energy in ecosystems. It is a prominent measure of the “scale” of human activities compared to natural processes (Costanza, 1992). Because human harvest of biomass is a major component of HANPP, it is also closely related to social metabolism (Haberl et al., 2019). The basic question of how much of the biosphere’s yearly biomass flows is used by humans was first posed in the 1970s (Whittaker and Likens, 1973). It took more than a decade until the first comprehensive answer to that question was given (Vitousek et al., 1986). This entry gives an overview of the research that has followed these seminal statements, and proceeds by discussing issues of defining HANPP (section 50.2), presenting some methodological basics (section 50.3), and giving an overview of the current knowledge on global HANPP (section 50.4). This is followed by a concluding section on interpretation and further research requirements (section 50.5).

Definition of HANPP

All measures of HANPP reflect human impacts on net primary production (NPP) and its availability in ecosystems under current land-use patterns. NPP is the total amount of biomass produced through photosynthesis, net of the plants’ own metabolic demands. However, different definitions of HANPP have been used, which has resulted in substantial differences across empirical results. This lack of standardization created the impression that it would be very difficult to assess HANPP with sufficient accuracy. Lacking standardization hampered the comparability of results and has resulted in criticisms (Davidson, 2000; Rojstaczer et al., 2001). Appreciation of the definitions of HANPP is key to better understanding these debates (Haberl et al., 2007). Vitousek et al. (1986) calculated HANPP using three definitions, each of which measures a different process or pattern. The first, most narrow definition only includes biomass directly used by humans as food or timber. The second, intermediate definition also includes the NPP of human-dominated ecosystems such as croplands. The third definition additionally considers the NPP lost due to human-induced changes in ecosystem productivity (e.g., as a result of ecosystem degradation). Wright (1990) proposed to define HANPP as the difference between the NPP available in the hypothetical, undisturbed ecosystems that would exist in the absence of land use and the amount of NPP actually available under current conditions to support heterotrophic food chains. He excluded activities such as logging and biomass burning in forests on the grounds that they would not result in a long-term reduction of productivity of the land for wild species if forests were allowed to regrow. There is ample evidence, however, that harvest and biomass burning are very important for forest ecology (Harmon et al., 1990, 1986), which is why other authors included timber harvest and human-induced fires in their definitions of HANPP (Haberl et al., 2014; Vitousek et al., 1986). Building on Wright’s definition, Haberl (1997) proposed a definition of HANPP that was later refined in subsequent work (deSouza and Malhi, 2018; Erb et al., 2009b; Haberl et al., 2014, 2007; Kastner et al., 2021; Krausmann et al., 2013; Zhang et al.,

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2021). This strand of work defines HANPP as the difference between the NPP of potential vegetation (NPPpot), that is, the vegetation that would occur without land use under current climate conditions, and the fraction of the NPP of the actually prevalent vegetation (NPPact) that remains in the ecosystem after human biomass harvest. The latter variable is denoted as NPPeco. This definition may be written in two different formulas, which both have the same meaning, but each represent a different perspective: ● Ecological perspective: HANPP = NPPpot – NPPeco ● Societal perspective: HANPP = HANPPluc + HANPPharv HANPPluc denotes the HANPP resulting from land conversion and land use and is defined as NPPpot – NPPact. HANPPharv denotes the HANPP resulting from biomass harvest and is defined as NPPact – NPPeco. Most current approaches define HANPPharv to include biomass of plant parts that are destroyed or killed during harvest, even if they remain on site (e.g., roots of harvested plants or biomass consumed in human-induced vegetation fires; Lauk and Erb, 2009). As with all indicators, differences in definitions substantially affect

Source:

the results of HANPP calculations (Haberl et al., 2007; Smil, 2013). The definition of HANPP shown in Figure 50.1 excludes NPP remaining in ecosystems that is available for wild-living organisms or accumulating over time; both may occur even in intensively managed ecosystems, such as managed forests, grasslands, or even cropland. It is also robust in time-series calculations. Land use sometimes reduces NPP (e.g., due to land degradation; Zika and Erb 2009), or even prevents it altogether (e.g., soil sealing), but technologies such as irrigation, fertilization, or use of improved crop varieties may also raise NPP over its natural potential, which is considered in this definition. In this case, HANPPluc or even total HANPP may assume negative values. Considering this possibility is important because such effects are significant and historically variable (Haberl et al., 2014; Kastner et al., 2021). For example, NPPact exceeds NPPpot in large parts of the intensively used agrarian landscapes in northwest Europe (Plutzar et al. 2016). Negative HANPP values can also be observed in irrigated dryland areas (Haberl et al., 2007). While raising yields may be considered as benign from an anthropocentric perspective, it often comes with substantial ecological detriments, such as nitrogen leach-

Redrawn after Haberl et al. (2014).

Figure 50.1 Definition of HANPP

Helmut Haberl, Karl-Heinz Erb, and Fridolin Krausmann

Human appropriation of net primary production (HANPP)  291

ing in humid zones and waterlogging or soil salinization in drylands (Singh, 2021). How HANPPharv should be defined is another important issue. For example, should biomass residues remaining on site and plowed into the soil be included or not? How can we appropriately deal with wood harvests in forests, where a stock accumulated over many decades is harvested? The answer usually depends on the purpose of the respective study, and slightly different definitions have been used in recent work (Haberl et al., 2014, 2007; Kastner et al., 2021; Krausmann et al., 2013). The HANPP indicators discussed so far refer to specific units of land (e.g., a national territory or any other spatially delimited unit). In the last one or two decades, another research question has gained prominence: Which impacts are related to the consumption of a defined population (e.g., a nation’s population)? In the HANPP context, this question was first posed by Imhoff et al. (2004) and later operationalized through the concept of “embodied HANPP,” abbreviated eHANPP (Erb et al., 2009c; Haberl et al., 2009; Kastner et al., 2011). eHANPP was defined as the HANPP caused – anywhere on the planet – by the consumption of a defined population, an approach that allows us to systematically study teleconnections in the land systems and analyze drivers of land-system change (Dorninger et al., 2021; Kastner et al., 2015; Roux et al., 2021; Saikku and Mattila, 2017).

50.3

Some basics of HANPP methodology

To calculate HANPP it is necessary to assess three properties: (1) NPPpot, the NPP of the vegetation that would be assumed to prevail in the absence of human land use but with current climate (potential vegetation); (2) NPPact, the NPP of the currently prevailing vegetation; and (3) HANPPharv, the human harvest of NPP. Different methods are available to estimate these three properties. Which one is most appropriate depends on the scope and purpose of the study. One of the strengths of HANPP is that it can be assessed in a spatially explicit way; that is, it is possible to produce maps of HANPP that localize the human impact on ecosystems. In this case, the three above-mentioned parameters must be calculated in a spatially explicit manner,

using geographic information systems methods (Haberl et al., 2007; Kastner et al., 2021). NPPpot is a hypothetical parameter that cannot, with the exception of wilderness areas, be observed or measured directly, but needs to be derived by models. Approaches of varying complexity have been suggested. Empirical models use correlations between measured climate variables and productivity from sites studies. The prominent Miami-Model (Lieth, 1975) used data collected during the International Biological Program (e.g., Cannell, 1982). State-of-the art, mechanistic dynamic vegetation models that simulate the exchange processes between plants and atmosphere, such as gross primary production and plant respiration, can also be used to derive NPPpot maps in time series by omitting land use as a driving parameter and allowing only climate parameters to drive the dynamics (e.g., Bondeau et al., 2007). HANPPharv is derived from agricultural and forestry statistics (e.g., from FAOSTAT; Food and Agriculture Organization [FAO], 2021), combined with material and energy flow methods in order to reflect not only reported primary harvest (e.g., grain, recovered wood), but also secondary biomass flows (e.g., recovered and unrecovered straw, wood felled but not removed from forests, biomass harvest in cities), and can be mapped using land-use maps that are consistent with census statistics and are of adequate thematic resolution. This is relatively straightforward for cropland, where data are reliable, but challenging for livestock grazing. Statistics on grazing are usually of poor quality, owing to the low economic value of grazing land and to existing ambiguities in definitions. For example, FAO statistics refer only to permanent pastures and do not report on occasional grazing. The amount of grazed biomass is usually not available and needs to be derived through modeling (Erb et al., 2017; Fetzel et al., 2018). While quantifying HANPPharv in forests is methodologically relatively straightforward (e.g., at the national scale), mapping is intricate because, in forestry, biomass stocks, not flows, are harvested. Many studies therefore do not map the location of harvest (e.g., of clear-cut areas) because HANPP results would appear implausible. Instead, harvest is allocated to the entire managed forest area (excluding

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untouched forests), for example, following the NPPact pattern (see next paragraph). This implicitly assumes that the entire managed forest area is harvested once during a full rotation cycle and that harvest pressure is the same in the entire forestry area under consideration (Erb et al., 2009b; Kastner et al., 2021). Various approaches are used to estimate NPPact. On cropland, NPPact can be extrapolated from data on crop harvest, using appropriate coefficients (Krausmann et al., 2013). NPPact on forestland and grassland is, due to the lack of adequate data, usually derived from NPPpot with factors that consider information on soil degradation and irrigation. For managed forests, most HANPP studies conducted so far have used the assumption that their NPP were equal to that of unmanaged forests (e.g., Kastner et al., 2021), although other approaches have also been implemented (O’Neill et al., 2007). The limited evidence available (Erb et al., 2016a; Noormets et al., 2015) suggests this approach on general terms: apparently, harvest results in younger stands in terms of succession, which are more productive, and reduces light competition. These boosting effects may be offset by the effects of forestry, such as the reduction of photosynthetically active tissue in forest ecosystems. However, effects of forest degradation are difficult to grasp (Arneth et al., 2021). For grazing land, proxies to account for degradation have been used (Zika and Erb, 2009). As data on belowground NPP are considerably more uncertain than those on aboveground NPP, many HANPP studies were restricted to the aboveground compartment, although some studies included belowground HANPP (e.g., Haberl et al., 2007). A focus on total NPP avoids ambiguities surrounding crops like roots and tubers (Smil, 2013) and is generally more meaningful because the fraction of NPP allocated to above- or belowground components differs strongly between various types of ecosystems. These differences exist for natural ecosystems under varying climate and soil conditions, and also depend on land use on managed lands. For example, converting a forest or natural grassland into a crop field strongly reduces the fraction or NPP allocated to belowground compartments. In any case, aboveground and belowground processes should be accounted for separately. More detailed information on HANPP

methods can be found elsewhere (Haberl et al., 2014; Kastner et al., 2021).

50.4

A brief history of global HANPP estimates

The first study of global HANPP (Whittaker and Likens, 1973) reported that humans harvested 1.6 Pg C/yr from terrestrial ecosystems as food and wood in the 1950s, which was 3 percent of their estimate of total global terrestrial NPP (54 PgC/yr) – a low value that resulted from the very narrow definition used. In 1986, global HANPP was estimated to amount to 3–39 percent, depending on the definition (Vitousek et al., 1986). Using largely the same data but a different definition, Wright (1990) estimated global HANPP at 24 percent. A probabilistic study using Vitousek et al.’s intermediate definition reported a large uncertainty range for global HANPP (Rojstaczer et al., 2001), a conclusion that attracted criticism (Field, 2001; Haberl et al., 2002). The consumption of HANPP was estimated to be 16–24 percent, again using a different definition (Imhoff et al., 2004). The first spatially explicit assessment of global HANPP (Haberl et al., 2007) used extensive data triangulation methods to generate consistent and spatially highly detailed data on land use (Erb et al., 2007) and biomass flows (Krausmann et al., 2008). This study found global total HANPP in the year 2000 to be 24 percent of NPPpot; aboveground HANPP was found to be substantially higher (almost 30 percent). A centennial time series showed that global HANPP doubled from 1900 to 2010 (Krausmann et al., 2013). A recent global mapping exercise quantified global terrestrial HANPP in 1900 at 13.5 percent and in 2010 at 20.7 percent (Kastner et al., 2021). Global HANPP results of various studies are shown in Table 50.1.

50.5 Meaning and significance of HANPP

HANPP has been interpreted as a measure of the physical size of the economy relative to the ecosystem in which it is embedded (Daly, 2007), a claim that attracted criticism (Davidson, 2000; Smil, 2013). HANPP calculations reveal how much of the trophic energy that would be available for wild-living animals and other heterotroph organisms in the absence of human activities is still in

Helmut Haberl, Karl-Heinz Erb, and Fridolin Krausmann

Human appropriation of net primary production (HANPP)  293 Table 50.1 Overview of estimates of global HANPP given by different authors Study

Reference time

HANPP absolute

HANPP relative

(Pg C/yr)

(%)a

Whittaker and Likens (1973)

1950s

1.6

3%

Vitousek et al. (1986) low

1970s

2.6

3%

Vitousek et al. (1986) intermediate

1970s

20.3

27%

Vitousek et al. (1986) high

1970s

29.5

39%

Wright (1990)

1970s–80s

17.7

24%

Rojstaczer et al. (2001)

1980s–90s

19.5±14

32% (10–55%)

Imhoff et al. (2004)

1995

Haberl et al. (2007)

2000

11.5 (8.0–14.8) 15.6

20% (14–26%) 24.0%

Krausmann et al. (2013)

1900

6.9

13.0%

Krausmann et al. (2013)

2010

14.8

25.0%

Kastner et al. (2021)

1910

7.1

13.5%

Kastner et al. (2021)

2010

12.8

20.7%

Note: The differences in definitions used in each study are discussed in the text. Biomass flows were converted from original data assuming that 1 kg dry matter biomass equals 0.5 kg carbon and has a gross calorific value of 18.5 MJ/ kg. a Per cent of actual or potential NPP. Estimates of NPPact and NPPpot also vary considerably; for example, Whittaker and Likens’s value of NPPact (54 PgC/yr) was much lower than Vitousek et al.’s estimate (66 PgC/yr).

place, thereby providing an indicator of the “human domination of ecosystems” (Vitousek et al., 1997). It can also be interpreted as a system-level indicator of land-use intensity (Erb et al., 2013) that can be mapped globally, even in centennial timescales (Kastner et al., 2021). Studies of global HANPP gained attention in the sustainability debate because HANPP was interpreted as an indicator for ecological limits to growth (Meadows et al., 1992; Sagoff, 1995). Others regarded it as a readily measurable indicator of an important planetary boundary (Running, 2012). This concept has lost appeal, however, because long HANPP time series (Kastner et al., 2021; Krausmann et al., 2013) showed that much potential exists to “decouple” HANPP from the growth of economic activity (GDP) and even from population. Economic growth is increasingly driven by energy sources other than biomass (Global Energy Assessment, 2012). Biomass harvest may grow as a result of agricultural intensification rather than through extension of farmed areas (Davidson, 2000; Kastner et al., 2021). When harvest indices are raised (i.e., the fraction of biomass allocated by the crop plant to the commercial product is increased), the amount of product per unit HANPPharv – that is, HANPP-efficiency – rises considerably (Krausmann et al., 2013). Feeding efficiencies in livestock production have also been increased, partly offsetting

trends toward higher animal-product consumption (Krausmann et al., 2009). The question of what HANPP and related indicators can contribute to understanding planetary boundaries in the context of land use is therefore more complex than often thought (Haberl and Erb, 2017). The HANPP approach has yielded a family of related assessments of human impacts on ecological patterns and processes. A major study (Erb et al., 2018) found that global land use reduces the amount of carbon in terrestrial vegetation by approximately one-half. A surprisingly small fraction of this effect – only a bit more than one-half – was found to be caused by deforestation, while land uses that did not change the dominant land cover (e.g., forestry or grazing on native grasslands) played a substantial role. Because land use affects biomass stocks more strongly than NPP, it results in a considerable acceleration of the carbon turnover rate of global vegetation – on average, one atom of carbon currently remains in vegetation for only seven years, half of the value found in the potential vegetation (Erb et al., 2016b). Major methodological progress has also been made in recent years, as new studies include large-scale uncertainty assessments (Erb et al., 2018; Erb et al., 2016a; Kastner et al., 2021). Assessing the option space for future agriculture, land use and bioenergy produc-

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tion has become another important application of the HANPP approach. It supported development of the Biomass Balance Model (BioBaM; Erb et al., 2009a) that was recently extended to the BioBaM-GHG model capable of calculating greenhouse gas emissions of future land use and agriculture scenarios (Kalt et al., 2021; Theurl et al., 2020). This modeling approach allowed delineating the option space for feeding the world in 2050 (Erb et al., 2016b) and to quantify associated greenhouse gas emissions (Theurl et al., 2020). Several studies used the BioBaM-GHG model to estimate future bioenergy potentials and their GHG costs per unit of bioenergy produced (Kalt et al., 2020). Such calculations also helped to constrain estimates of future sustainable potentials for removing CO2 from the atmosphere with Bioenergy with Carbon Capture and Sequestration (BECCS) technologies (Creutzig et al., 2021). Overall, these modeling approaches provide more nuanced assessments of the planetary boundaries related to land (Haberl and Erb, 2017) and confirm that demand-side options (e.g., dietary change) are key to climate change mitigation in the global land system. HANPP can also serve as an indicator for societal pressures on biodiversity. The “species-energy” hypothesis holds that the availability of trophic energy in ecosystems is one of several determinants of biodiversity (Gaston, 2000). Because HANPP reduces energy availability in ecosystems, it can help gauging global human impacts on species richness (Wright, 1990). Empirical research into energy-species relationships in human-dominated landscapes so far supported the notion that HANPP may co-determine species richness (e.g., Lorel et al., 2019). Spatially explicit HANPP information was also used to map and quantify the impact of land-use intensity on biodiversity (Semenchuk et al. 2022), a still severely under-researched topic (Dullinger et al., 2021). Studies of embodied HANPP have, among others, helped to demonstrate the growing dependence of the European Union on imports from other world regions, notably Latin America (Kastner et al., 2015). A recent study of embodied HANPP contradicted the notion that international trade would contribute to land sparing, allegedly because it would help shifting agricultural production to

the regions with the most efficient land-use systems (Roux et al., 2021). We conclude that the analysis of socio-economic drivers of HANPP and its ecological impacts remains highly relevant for sustainability science and ecological economics.

Acknowledgments

We acknowledge support by the Austrian Science Funds (FWF, www​.fwf​.ac​.at), projects P16692, P20812, and P29130, and by the Volante project funded by the European Union’s Framework Programme 7 (FP7), grant no. 265104. Helmut Haberl, Karl-Heinz Erb, and Fridolin Krausmann

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appropriation of net primary production doubled in the 20th century. PNAS 110, 10324–9. Krausmann, F., Erb, K.-H., Gingrich, S., Lauk, C., Haberl, H., 2008. Global patterns of socioeconomic biomass flows in the year 2000: a comprehensive assessment of supply, consumption and constraints. Ecological Economics 65, 471–87. Krausmann, F., Haberl, H., Erb, K.-H., Wiesinger, M., Gaube, V., Gingrich, S., 2009. What determines geographical patterns of the global human appropriation of net primary production? Journal of Land Use Science 4, 15–33. Lauk, C., Erb, K.-H., 2009. Biomass consumed in anthropogenic vegetation fires: global patterns and processes. Ecological Economics 69, 301–09. Lieth, H., 1975. Modeling the primary productivity of the world. In Lieth, H., Whittaker, R.H. (eds.), Primary productivity of the biosphere, ecological studies 14. Springer, 237–63. Lorel, C., Plutzar, C., Erb, K.-H., Mouchet, M., 2019. Linking the human appropriation of net primary productivity-based indicators, input cost and high nature value to the dimensions of land-use intensity across French agricultural landscapes. Agriculture, Ecosystems & Environment 283, 106565. Meadows, D.H., Meadows, D.L., Randers, J., 1992. Beyond the limits: global collapse or a sustainable future. Earthscan. Noormets, A., Epron, D., Domec, J.C., McNulty, S.G., Fox, T., Sun, G., King, J.S., 2015. Effects of forest management on productivity and carbon sequestration: a review and hypothesis. Forest Ecology and Management, Carbon, Water and Nutrient Cycling in Managed Forests 355, 124–40. O’Neill, D.W., Tyedmers, P.H., Beazley, K.F., 2007. Human appropriation of net primary production (HANPP) in Nova Scotia, Canada. Regional Environmental Change 7, 1–14. Plutzar, C., Kroisleitner, C., Haberl, H., Fetzel, T., Bulgheroni, C., Beringer, T., Hostert, P., Kastner, T., Kuemmerle, T., Lauk, C., Levers, C., Lindner, M., Moser, D., Müller, D., Niedertscheider, M., Paracchini, M.L., Schaphoff, S., Verburg, P.H., Verkerk, P.J., Erb, K.-H., 2016. Changes in the spatial patterns of human appropriation of net primary production (HANPP) in Europe 1990–2006. Regional Environmental Change 16, 1225–38. Rojstaczer, S., Sterling, S.M., Moore, N.J., 2001. Human appropriation of photosynthesis products. Science 294, 2549–52. Roux, N., Kastner, T., Erb, K.-H., Haberl, H., 2021. Does agricultural trade reduce pressure on land ecosystems? Decomposing drivers of the embodied human appropriation of net

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51. The human ecological footprint Setting the stage

The first question of human ecology—and therefore ecological economics—should be ‘Just how much of Earth’s surface is dedicated to supporting just me in the style to which I am accustomed?’ Everyone has one. Without it no one could live, though few people are even aware of its existence. I am not referring here to such vital organs as the heart or brain, but rather to something somewhat less obvious—our ecological footprints (without which we would have neither hearts nor brains; we couldn’t exist). What, exactly, do we mean by ecological footprint (EF)? The short answer is that the EF is an area-based estimate of human energy and material demands on the ecosphere. More formally: ‘The EF of any specified population is the total area of productive land and water ecosystems required, on a continuous basis, to produce the bio-resources that the population consumes and to assimilate its (carbon) wastes, wherever on Earth the relevant land/ water ecosystem are located’ (Rees, 2013, emphasis in original). In practice, a population’s EF includes all the cropland, grazing land, forested land, fishing grounds, and carbon sink ecosystems, together with the built-up (usually urban) land areas, required to provide the biomaterial demands (e.g. food and fibre) of the population. This aggregate area represents the study population’s appropriation from Earth’s total biocapacity. Eco-footprint analysis (EFA) is based on several undeniable facts. First, every human economic activity—indeed, every activity of any kind—necessarily involves the use/ dissipation of energy and the transformation/ consumption of material resources; second, a measurable portion of those resources (e.g. plant-based food and fibre products, livestock and fish products, timber and other forest products) is produced by terrestrial or aquatic ecosystems; third, we can trace the flow of resources from their origins in land/water ecosystems, through the economy and back to the ecosphere (all material flows return as waste); fourth, we can convert many of these flows to corresponding productive or assimilative ecosystem areas. Assuming ade-

quate data are available—preferably on an annualized basis—EF estimates can be made for any population, from single individuals to the entire human family (we can also make EF estimates for specific economic sectors or individual products). The size of a population’s EF depends on four factors: size of the population; average material standard of living (consumption); the productivity of relevant ecosystems; and the technological efficiency of resource harvesting, processing, and manufacturing. At the national level, EF estimates are based on each country’s final consumption of energy, bulk commodities, and manufactured goods, where consumption is defined as: domestic production plus imports minus exports. A major strength of EFA is that, once we have estimated a nation’s EF, we can compare it to that country’s biocapacity (its domestic productive ecosystem area). EFA is the only aggregate sustainability assessment tool that enables comparisons of humanity’s demand for biocapacity with the available supply. Since different ecosystems and management practices (e.g. fertilizer, irrigation) result in differing ecosystem productivities, analysts generally convert EF and biocapacity estimates to their equivalents in global average hectares (gha). For example, suppose country A has arable land that is, on average, twice as productive as global average cropland. In estimating A’s domestic biocapacity, EF analysts would treat x actual hectares of A’s cropland as equivalent to 2 x gha. Normalizing data so that all numbers are in gha facilitates fair comparisons of the EFs and biocapacities of nations.

A note on EF data sources

The Global Footprint Network (GFN) is the major nongovernmental organization (NGO) engaged in EFA. GFN and its allied Footprint Data Foundation maintain national EF accounts for some 200 countries and territories (GFN, 2021a) based on United Nations or UN-affiliated data sets, including those published by the Food and Agriculture Organization, UN Commodity Trade Statistics Database, the UN Statistics Division, and the International Energy Agency (details of current methods, data, and results are available by following the links at www​.footprintnetwork​.org/​).

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How large is the human footprint and does it matter?

The total area of productive land and aquatic ecosystems on Earth is about 12.1 billion hectares (ice caps and deserts are not included, nor is most of the open ocean which is effectively marine desert). This seems like an enormous area until compared to the human EF of 20.9 billion hectares (2017 data from GFN, 2021a). Eco-footprint data dramatically underscore the extent of humanity’s current (un)sustainability conundrum. The human enterprise is exploiting the ecosphere as if the Earth were 73 per cent larger than it is. To put it another way, humanity has ‘overshot’ Earth’s biocapacity by 73 per cent—we are running a massive ecological deficit.

Explaining overshoot

How can this be? Overshoot is possible only because of the tremendous stocks of so-called ‘natural capital’ that have accumulated in various ecosystems and physical systems and because of the enormous assimilative capacity of the ecosphere. Fertile soils, rainforests, fish stocks, and so on take centuries to accumulate to mature capacity; less renewable and non-renewable ‘assets’ such as ground water and fossil fuels build up only on a geological time scale. The problem is that, at present, the human enterprise (population and economy) maintains itself and grows, not only by using up the annual production of its supportive ecosystems (so-called ‘natural income’), but also by consuming a portion of the essential capital stocks. The enormous (often toxic) waste load generated by economic (over) production/consumption spews back into the ecosphere, further impairing production. The world is fixated on a single issue, climate change, but EFA more fully documents modern civilization’s true eco-predicament. Global warming as well as plunging biodiversity, overfishing, tropical deforestation, soil/land degradation, the pollution of air, land, and waters, among others, are all co-symptoms of overshoot. Overshoot is clearly the meta-problem. None of its major symptoms can be cured in isolation—the present focus on climate change will not only fail to reverse climate change, but acceptable approaches (massive investment in energyand material-intensive techno-fixes) will worsen overshoot.

In 2021, Earth Overshoot Day occurred on July 29. Overshoot Day marks the annual date by which “humanity has exhausted nature’s budget for the year” (GFN, 2021b). For the rest of the year, we make up for overshoot by drawing down resource stocks, accumulating carbon dioxide in the atmosphere, and dissipating other wastes throughout the ecosphere (i.e., by literally depleting and polluting the biophysical basis of our own existence). Untreated, overshoot is, by definition, a fatal condition.

Sustainability implications of EFA

It should be obvious from even the above that, in functional terms, population EFs represent ecologically exclusive areas. Biocapacity required by one person or nation is not available for use by another. On a finite planet in overshoot, this means that every individual and all populations are in competition for the declining biocapacity of Earth. And it’s not a truly fair competition. In a world characterized by an increasingly global economy and integrated markets, those with the most money win. People in high-income countries enjoy the material benefits implied by EFs of 4–6 gha, as is typical of Western European countries (although Luxembourg comes in at 12 gha). Canadians, Americans, and Australians have average EFs in the 7–8 or more gha range. By contrast, millions of people in poorer African (e.g. Congo, Zambia, Kenya) and Asian (e.g. Pakistan, Afghanistan, Bangladesh) countries get by on the goods and services provided by less than 1 gha. In high-income countries, the demand for carbon sink ecosystems comprises well over half the national EF, but only a small fraction of poor country EFs. EFA clearly reflects both the lock-step relationship between gross domestic product per capita and fossil energy consumption, as well as the egregious and increasing income inequality among the world’s nations. There is another problem. Simple division of the world’s 12.1 hectares of productive ecosystems among the world’s 7.5 billion people in 2017 would allocate ~1.6 gha to each member of the human family (without considering the needs of non-human species competing for the same biocapacity). However, with a global EF of 20.9 billion ha, the average per capita human EF was ~2.8 gha. (The difference illustrates average William E. Rees

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individual overshoot.) In short, Europeans’ EFs are typically twice the world average and more than three times what they would receive if biocapacity were equally distributed; North Americans’ EFs are 2.9 times the world average and fully five times larger than their hypothetical proportional share. Put another way, we would need four additional Earth-like planets to support just the present world population at North American material standards. By contrast, denizens of the world’s poorest countries must subsist on the goods and services provided by as little as half their proportional share of biocapacity. Again, EFA underscores the disparity between rich and poor and graphically shows how money wealth determines people’s access to the essential goods and services of nature.

On sustainability with justice

Clearly, EFA can help to estimate the scope of structural change needed if the world is to survive the ecological crisis. To begin, we must eliminate global overshoot. ‘One-planet living’ implies that, with today’s 7.5 billion people, average EFs cannot exceed 1.6 gha. The data above suggest that, to achieve a just sustainability, global energy and material consumption should be reduced by 40–50 per cent and by up to 80–90 per cent in high-end consumer countries like Canada and the US. This would allow for much-needed consumption increases by peoples currently living on less than their equitable allocation. Keep in mind, however, that these are already dated (2017) estimates—the 2022 population is approaching 8 billion at the rate of 80 million additional people per year, and biocapacity is being eroded daily by everything from accelerating climate change to severe over-exploitation. Every moment of delay makes the transition more difficult. Reducing global consumption by 50 per cent or more may seem extreme but is: a) necessary to eliminate overshoot and its symptoms—catastrophic climate change, ecosystems collapse, gross pollution, and so on; and b) fully consistent with several other assessments, including one by the Business Council for Sustainable Development (BCSD, 1993) as long ago as 1993, and others up to the present day (e.g. Akenji et al., 2021; Rees, 2020a). Some would argue that such reductions are also technologically achievable William E. Rees

(along with appropriate life-style changes) and would actually improve well-being (von Weizsäcker et al., 2009). Certainly everyone would also benefit from the reduced pressure on biocapacity and life-support services that would flow from lower human numbers. Indeed, to ignore ‘the population problem’ is untenable: to achieve sustainability on just one Earth without attention to population reduction would require that most people become/remain poor. Bottom line: EFA shows unambiguously that the world community can only gain from a vigorous de-growth movement based on the synergism among technological gains, reduced economic production/consumption, and non-coercive population reduction policies.

Overshoot, eco-deficits, and the unsustainability of nations

EFA is closely related to the concept of carrying capacity (CC). In wildlife ecology, CC is defined as the average maximum population of a species that a particular habitat can support indefinitely. Thus, while CC asks how large a population a given area can support, EFA inverts the CC ratio by asking how large an area is needed to support a given population. It might seem that if every individual requires a certain minimum productive ecosystem area, there must be an upper limit to the number of people any country can support; this is its domestic CC. ‘Nonsense,’ economists protest, ‘why should local populations be limited by local biocapacity if they can trade for essential goods? Don’t globalization and trade increase every country’s domestic CC? And technology is constantly improving efficiency or substituting for natural resources. Surely the open exchange of goods and ideas in an integrated global market frees human populations everywhere from local ecological constraints’. Not so fast! While trade seems to inflate national CCs, it neither decreases local demand nor increases global limits; what it does do is shuffle bits of biocapacity around. Trade enables nations to import biocapacity embodied in trade goods from distant ‘elsewheres’ while simultaneously isolating their people spatially and psychologically from the ecosystems that support them. As a result, importing nations can vastly exceed their

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domestic CC while drawing down exporting countries’ and the global commons’ remaining stocks of natural capital. Globalization/ trade also exposes Earth’s shrinking stocks of biocapacity to the largest possible (and growing) population of ever-wealthier consumers. Meanwhile, by making goods cheaper and more wildly available, efficiency gains have historically increased consumption. Bottom line: globalization/trade enables nations to run massive eco-deficits and increases the rate at which humans over-exploit and degrade the Earth. This situation impacts long-term sustainability in at least three ways: 1) the swollen populations of importing countries become dependent on potentially unreliable supplies of eco-goods and services produced half a planet away; 2) the export-driven erosion of exporting countries’ domestic biocapacity compromises their future options, should circumstances change; and 3) the inflated global population and increased competitive demand for shrinking biocapacity both reduces ecological resilience and increases geopolitical tension. Ominously, almost 150 of the jurisdictions (three-fourths of the world’s nations) for which GFN maintains EF data are currently running ecological deficits (GFN, 2021c). Diminutive, densely populated Singapore is exceptional in that the country’s EF exceeds its domestic biocapacity by over 10 000 per cent, but most Western European nations have deficits in the 500 per cent to 1000 per cent range; Japan’s eco-deficit is almost 700 per cent. The majority of the 150 deficit countries have EFs at least twice their domestic biocapacities. All such nations live largely on imported biocapacity, a dangerously unsustainable and geopolitically unstable situation in a world of accelerating change. Trade is totally fuelled by fossil energy. Meanwhile, the world is striving to decarbonize to avoid climate catastrophe and we are simultaneously depleting economic sources of oil. It is quite clear that renewable wind and solar electricity (recently only 4 per cent of global primary energy supply) will not be able to substitute for oil, in any climate-relevant time-frame (Seibert and Rees, 2021). Global supply chains are already breaking. These realities herald the end of the era of expanding trade and global growth. The question is:

How will trade-dependent nations, home to billions of inhabitants who could not survive on domestic CC, respond to soaring energy costs, food scarcity, loss of access to other essential resources, shrinking economies, civil unrest, and the prospects of millions of desperate transnational migrants?

Overshoot—an existential threat to cities

Cities are all eco-deficit. EF studies of cities have long revealed that modern techno-industrial (MTI) cities have EFs hundreds or even a thousand times larger than their political or built-up areas, depending on population densities and their citizens’ lifestyles (e.g. Folke et al., 1997; Warren-Rhodes and Koenig, 2001; Rees 2020b). The reason is obvious. Cities, per se, are incomplete ecosystems, merely consumptive nodes in the widely dispersed human ecosystem; the productive complement—99 per cent+ of the total area—comprises extra-urban land/ waterscapes scattered all over the planet (Box 51.1). In short, most cities have essentially no domestic biocapacity—they rely entirely on imports of resources from, and exports of waste to, their globally dispersed hinterlands. BOX 51.1 THE STRUCTURE AND FUNCTION OF ECOSYSTEMS Cities are not functionally complete ecosystems. A complete ecosystem is a self-organizing or ‘autopoietic’ living system comprising numerous mutually interdependent species and a well-defined trophic (feeding) structure. Complete ecosystems include: 1) producer organisms (mostly green plants); 2) macro-consumers (mostly multi-cellular animals, including humans); and 3) micro-consumers (bacteria and fungi). Green plants are able to self-produce using sunlight (photosynthesis) to assemble biomass from carbon dioxide, water, and trace nutrients extracted from the soil. Macro-consumers self-produce using energy and material obtained by consuming plants or other animals; and micro-consumers self-produce with the energy/matter gained by decomposing the bodies of both plants and animals and returning surplus nutrients and organic matter to the soil. Energy thus cascades unidirectionally and irreversibly

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302  Elgar encyclopedia of ecological economics through ecosystems, while decomposition enables the material cycle to repeat continuously. In short, unlike cities, which lack adequate complements of producers and micro-consumers, complete ecosystems are exquisitely complex quasi-independent systems. Their full complement of species functions in such a way that the entire complex can produce and maintain itself ‘far-from-equilibrium’ indefinitely by dissipating solar energy. The resultant waste heat is radiated off the planet, increasing the entropy of the universe.

This is another way of saying that modern cities are the product of fossil fuels and remain dependent on fossil-fuelled air, highway, and marine transportation for their continued existence. So, again, as MTI society attempts to decarbonize, cities will confront a conundrum largely of our own making. It is improbable that electricity can replace fossil fuels in transportation in the coming decade. Therefore, to survive, our urbanizing civilization may choose to remain substantially reliant on oil—also coal and natural gas—at least while supplies last. This seems to be MTI society’s default position. Atmospheric CO2 and other greenhouse gas concentrations will increase, and global warming will likely exceed the 2°C upper mean global warming limit set by the Paris Agreement of 2015. Indeed, our current trajectory implies 3–4°C warming, which spells widespread climate disaster by late century—longer heat waves and droughts, accelerating desertification, melting permafrost, methane releases, water shortages, failing agriculture, possible famine, mass migrations, civil unrest, rising sea levels, the flooding (and eventual loss) of many coastal cities. Geopolitical conflict is almost inevitable. Urban life would become untenable, especially in the poorer, more vulnerable parts of the world. In the worst case, the world risks crossing tipping points (irreversible positive feedbacks) leading to runaway climate change and the end of anything passing for civilization. On the other hand, if the world attempts to avoid climate disaster through vastly increased investment in green energy or accelerated phasing out of fossil fuels, we could face serious energy shortages and shrinking economies even as global population and demand for everything increases. Reduced goods production, declining incomes, rising William E. Rees

inequality, widespread unemployment, falling agricultural output, broken international supply lines, failing inter-city transportation, local famines, and so on are again a recipe for the implosion of urban civilization. In short, whichever way we turn, large cities and megacities are likely to be cut off from their extended EFs and will not survive the spectre of compound eco-crises and geopolitical chaos that haunts the human future (Rees, 2020b). In the absence of a planned contraction, the chaotic collapse of urban civilization is a looming possibility.

Brickbats imagined and real

Are things really that bad? EFA has its share of critics who say ‘no!’ Most argue that the method doesn’t account for technological change, that it condemns growth or is biased against globalization. These charges are invalid. First, a population footprint is a static snapshot of reality at the time of analysis. If technological change has any effect on that population’s EF, this increase/decrease will be reflected in any subsequent ‘snapshot’. Second, all the EF provides is a quantitative estimate of appropriated ecosystem area. This estimate says nothing whatever about growth or globalization. Of course, using EFA results to question growth and undercut globalization is a perfectly legitimate use of the tool. It seems that many critics of EFA—mostly economists—are more concerned about the more obvious interpretations of the results than they are about flaws in the method. That said, EFA is hardly perfect. A significant weakness is an inherent tendency to underestimate human ecological impacts. EFA does not account for most diffuse forms of pollution, for example. Another major problem is that, while EFA can account for the total area of arable land, forest, grassland, and so on, effectively appropriated by a study population, it can say nothing about whether the relevant ecosystems are being ‘harvested’ sustainably. Agricultural soils, for example, are typically being degraded 10–40 times their regeneration rate. The implications of these flaws are disquieting. Applying ‘corrective factors’ to existing methods for EFA would significantly increase estimates of most population EFs. In short, global overshoot is worse, and implications are considerably more dire than what is outlined above.

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Epilogue – can EFA help ‘reconnect’ humanity to the ecosphere?

If overshoot is such a threat to the survival of urban civilization, why are most of the world’s people ignorant of the concept and its implications? Part of the answer is that the denizens of MTI societies rarely think of themselves as ecological beings. The prevailing neoliberal economic narrative considers the human enterprise to be separate from, and independent of, ‘the environment’; we assume human ingenuity (technology) can substitute for natural capital. Neoliberal economic models therefore contain no useful information about the complex structure and function of the ecosystems, or even the social systems, with which the growing economy interacts in the real world; we act as if H. sapiens were not subject to the natural laws that constrain other species (Rees, 2020a). By contrast, EFA, like other tools in ecological economics, recognizes that the human enterprise is a fully contained, dependent subsystem of the non-growing ecosphere. From this perspective, human exceptionalism is sheer illusion (Rees, 2020a). Ecologists recognize H. sapiens as a macro-consumer, a member of one of the three major ‘trophic’ components of every complete ecosystem (see Box 51.1). Despite the alienating influences of technology, globalization, and urbanization, humans are enormously significant ecological entities. In fact, EFA suggests we are actually the dominant macro-consumer in ecosystems all over the planet. Ironically then, while economists debate the degree to which the economy can ‘dematerialize’ from energy/material use or ‘decouple’ from the ecosphere (Kemp-Benedict, 2017), EFA shows unambiguously that human demands on nature steadily increase with population and income growth. And to dramatic effect—the inflated human enterprise has competitively displaced hundreds of other species from their habitats and food niches. From less than 1 percent 10 000 years ago, humans now comprise 36 per cent, and our domestic livestock another 60 per cent, of the planet’s expanded mammalian biomass, compared to only 4 per cent for all wild species combined (Bar-On et al., 2018). Consistent with EFA, these data suggest that H. sapiens may well be the most rapa-

ciously ‘successful’ carnivore and herbivore ever to walk the Earth. The problem is that humans’ evolutionary success, like that of any ill-adapted parasite, is undermining the structural integrity and function of its host, the living ecosphere. Yet there is room for hope; better than any other sustainability indicator, EFA graphically describes and quantifies the extent of human overshoot, the growing eco-deficits of most of the world’s nations, and the vulnerability of major cities to accelerating global change. To the extent that EFA can help raise ‘overshoot-as-fatal-condition’ to consciousness, it may help break the world’s tenacious allegiance to the perpetual growth ethic and become a catalyst for corrective action. In short, EFA may give H. sapiens a chance to redeem itself in the never-ending struggle up the evolutionary ladder. William E. Rees

References

Akenji, L., M. Bengtsson, V. Toivio, M. Lettenmeier. 2021. 1.5-Degree Lifestyles: Towards a Fair Consumption Space for All. Hot or Cool Institute, Berlin. https://​hotorcool​ .org/​1​-5​-degree​-lifestyles​-report/​ Bar-On, Y.M., R. Phillips, R. Milo. 2018. The biomass distribution on Earth. PNAS 21(May), 201711842. http://​www​.pnas​.org/​content/​early/​ 2018/​05/​15/​1711842115 Business Council for Sustainable Development (BCSD). 1993. Getting Eco-Efficient. Report of the BCSD First Antwerp Eco-Efficiency Workshop, November 1993. Geneva: BCSD. Folke, C., A.-M. Jansson, J. Larsson, R. Costanza. 1997. Ecosystem appropriation by cities. Ambio 26, 167–72. Global Footprint Network (GFN). 2021a. Open Data Platform–Country Trends. https://​data​ .footprintnetwork​.org/​?​_ga​=​2​.150429430​ .1954153549​.1635118484​-885488540​ .1614640094​#/​countryTrends​?cn​=​5001​&​type​=​ BCtot​,EFCtot GFN. 2021b. Open Data Platform. https://​data​ .footprintnetwork​.org/​#/​ GFN 2021c. Earth Overshoot Day. https://​ www​.footprintnetwork​.org/​our​-work/​earth​ -overshoot​-day/​ Kemp-Benedict, E. 2017. Dematerialization, Decoupling, and Productivity Change. Working Paper 1709, Post-Keynesian Economics Study Group. http://​www​.postkeynesian​.net/​ downloads/​working​-papers/​PKWP1709​.pdf Rees, W.E. 2013. Ecological Footprint, Concept of. In Levin, S.A. (ed.), Encyclopedia of

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304  Elgar encyclopedia of ecological economics Biodiversity (2nd ed., vol. 2). Waltham, MA: Renewable Energy Transition. Energies 14(15), Academic Press (pp. 701–13). 4508. https://​doi​.org/​10​.3390/​en14154508 Rees, W.E. 2020a. Ecological Economics von Weizsäcker, E., K. Hargroves, M. Smith, for Humanity’s Plague Phase. Ecological C. Desha, P. Stasinopoulos. 2009. Factor 5: Economics 169(March), 106919. https://​ doi​ Transforming the Global Economy through .org/​10​.1016/​j​.ecolecon​.2019​.106519 80% Increase in Resource Productivity. Rees, W.E. 2020b. MegaCities at Risk: The Droemer, Germany: Earthscan. Climate–Energy Conundrum. In Sorensen, Warren-Rhodes, K., A. Koenig. 2001. Ecosystem A., Labbe, D. (eds.), Handbook of Megacities Appropriation by Hong Kong and Its and Megacity Regions. Cheltenham, UK, and Implications for Sustainable Development. Northampton, MA: Edward Elgar Publishing Ecological Economics 39, 347–59. (pp. 292–308). Seibert, M., W.E Rees. 2021. Through the Eye of a Needle: An Eco-Heterodox Perspective on the

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52. Incommensurable values Incommensurable values pose a fundamental challenge for any approach to environmental management that assumes all environmental and non-environmental impacts and values can be expressed in monetary terms. Many orthodox approaches make this assumption, including cost–benefit analysis (CBA), payments for ecosystem services, and offset markets. The underlying idea behind claims about incommensurable values is the view that expressing all environmental values or impacts in monetary terms is at best unhelpful, usually misleading, and perhaps in some cases impossible. It is worth clarifying terminology. Alternative options or choices are incomparable when they cannot be ranked on an ordinal scale: if two alternatives x and y are incomparable, it is not true that ‘x is better than y’ or ‘y is better than x’ or ‘x is equal in value to y’. Alternatives are incommensurable when they cannot be precisely measured along some common cardinal scale of units of value: it is not possible to express how much better one alternative is than another, even though the alternatives may be comparable. A special case of commensurability is monetary commensurability, where the alternatives can be ranked along a monetary scale. Hence, comparability is a necessary but not sufficient condition for commensurability, and commensurability is a necessary but not sufficient condition for monetary commensurability. Since orthodox approaches to environmental management generally require monetary commensurability, it will be our focus in what follows. We examine three arguments used to justify monetary commensurability, that is, the view that the value of environmental impacts can always be meaningfully understood in monetary terms.

Argument 1: implicit valuation

In daily life people routinely accept an increased risk of death, in return for a monetary gain or saving. Person A buys an older used car, manufactured when safety standards were lower; B buys a new bike, but not the helmet, which is recommended for use

with it; C makes a detour down a dangerous road to save money at a shop there. These choices reveal an implicit monetary valuation of increased risk of death. The argument concludes by claiming that, if people implicitly put a money value on increased risk of death, they should be able and willing to value environmental impacts in monetary terms too. This argument is made frequently by economists (e.g. Frank, 2001), but it involves at least two problematic steps. The first is the assumption that these choices reveal a preference for ‘increased risk of death and more money’ over the status quo. The claim that choice reveals preference is famously controversial in economics (Hausman and McPherson, 2009). An alternative may be chosen not because it is preferred – the chooser is unable to rank the alternatives – but because it is better to choose something rather than nothing (Sen, 1997). In general, preferences cannot be inferred from choices without information about the chooser’s beliefs. When A buys an older used car, A may be unaware of the lower safety standards prevailing at the time of its manufacture. When C makes a detour down a dangerous road, C may do so just out of habit, or because C believes she is a safer driver than the average. Based on observing behaviour alone, it is not possible to infer a willingness to tolerate increased risk of death for a small monetary gain. Second, even supposing this inference can be drawn, we cannot conclude that willingness to ‘trade-off’ money for risk of death extends to other contexts. For example, it does not follow that there exists some level of financial compensation that would make A, B, or C agree to the siting of a nuclear waste storage facility near their home. These people are not being inconsistent if they refuse any compensation offer, because the mortality and morbidity risks posed by a nuclear waste storage facility may be different, along multiple qualitative and quantitative dimensions, to the money-risk trade-offs made by A, B, and C in their private consumption choices. The same point applies more strongly still when the environmental impacts being valued do not pose mortality and morbidity risks to humans, but involve other losses, such as the destruction of a unique wilderness area. Rather than compare such a loss to the willingness to pay for mortality risk reduction in private consumption choices, a better

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analogy might be with, say, the loss of a close friend. It is far from obvious that most people are able to put a money value on such a loss.

Argument 2: money is a neutral scale on which to measure and compare options

In the economic theory underlying techniques such as CBA, a common unit of measurement (‘the numeraire’) to compare diverse impacts is required, but it need not be money. As a CBA textbook states: ‘There is absolutely no need for money to be the numeraire (i.e. the unit of account) in such valuations. It could equally well be bushels of corn but money is convenient’ (Layard and Glaister, 1994: 2). Against this assumption that money is a convenient, neutral numeraire, it is clearly not neutral in at least one respect: the marginal value of money differs across individuals. One reason for this is very familiar and widely accepted by economists – the diminishing marginal value of wealth. Thus, when a rich person values an environmental impact at $x, this is less significant than an identical $x valuation from a poor person. The theory of CBA and related methodologies recommends the use of distributional weights to address this discrepancy. But there is another reason why the marginal value of money differs across individuals. Put simply, some people value money less because they place less value on the goods and services that can be bought or sold for money. For example, Brekke (1997) outlines how an ‘environmentalist’ may value money relatively less than a ‘materialist’, because the environmentalist places relatively less value on the goods and services that can be exchanged for money. Brekke’s argument merely assumes that there are some valued goods and services that are outside the market, broadly defined: they cannot be bought or sold for money. If distributional weights are required to correct for differing marginal valuations of money reflecting differences in wealth, it appears they are also required for differing marginal valuations of money reflecting different preferences for market goods and services. It is very hard to see how these weightings could be determined. A broader objection to the claimed neutrality of a money numeraire follows from the Jonathan Aldred

social meanings attached to monetary valuation. For example, there is strong evidence that people sometimes object to the offer of financial compensation for the imposition of some environmental damage or risk. The reason is that the compensation offer may be seen as an attempt to control or manipulate their behaviour (Frey and Jegen, 2001; Frey and Stutzer, 2008; Grant, 2012). It appears that these people do not interpret money as a neutral value against which environmental losses or risks can be readily compared. Research also suggests that ‘in-kind’ or ‘like-for-like’ compensation may be preferred to its monetary equivalent (Claro, 2007; Mansfield et al., 2002). Another relevant strand of research draws on interviews with respondents to contingent valuation surveys (or similar techniques for eliciting monetary valuations of environmental impacts). Some interviewees believe that stating a monetary valuation for some environmental impacts would be morally wrong or mistaken because it falsely implies that money is a valid substitute for the environmental asset or benefit. The refusal to express the value of something in monetary terms may form an essential part of a person’s ethical commitment towards that thing: that is, the refusal to value in monetary terms itself constitutes a particular value relationship or commitment to something (termed ‘constitutive incommensurability’ by O’Neill, 2007: 21–35). When people say that some environmental feature is ‘priceless’ or ‘irreplaceable’ or ‘unique’, they mean exactly that. They do not mean ‘highly valued’. In response to claims such as these, defenders of the monetary valuation of environmental impacts often argue that the money values entered in the economic analysis have no wider social meaning. Hansson (2007: 177) offers a clear statement of this view, in the context of valuing human life: Human lives do not have a monetary price in the common sense of the word. A cost–benefit analyst who assigns a monetary value to the loss of a human life does not thereby imply that someone can buy another person, or has the right to kill her, for that price. Essentially, lives and money are incommensurable, and the values of lives included in a CBA are for calculation purposes only.

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(Hansson also introduces ‘calculation values’ when discussing environmental damage, referring to them as being derived via contingent valuation surveys.) Hansson is clearly right to state that the calculation values associated with the loss of a human life do not have meanings such as the right to kill for that price. But what, exactly, is the meaning of these calculation values? This will depend on the context: how the values are derived, and the meaning attached to them by the economic analysis in which they are entered. However, in almost all cases, when some environmental impact is given a money value as part of a CBA, this implies that other impacts with a higher valuation have greater weight. If a new road project causes the destruction of a nature reserve valued at $v, but brings economic development benefits valued at $b, with b > v, then the CBA may recommend building the road, on the grounds that the economic benefits are worth more than the nature reserve. It is exactly this meaning, this statement of comparative worth, that is rejected by critics of CBA claiming that environmental impacts are incommensurable with money. The objection here is not a new one. It takes us back to Adam Smith’s diamond-water paradox. The CBA’s calculation values are in effect Smith’s ‘values in exchange’: b > v implies that we are willing to exchange the nature reserve for the road. As Smith explained, exchange values are often unrelated to the fundamental use (or more generally, well-being) values of the objects in question. (And the neoclassical argument that use and exchange values are equal ‘at the margin’ does not appear relevant to public policy decisions involving non-marginal environmental changes.) In sum, money values within CBA function as exchange values, and it is this social meaning which critics oppose.

Argument 3: for rational decision-making, trade-offs are unavoidable

Economists emphasize that choices have opportunity costs: any choice involves giving up something else of value. This is almost always true. However, it does not imply that we must compare the alternatives on a monetary scale (Lukes, 1996). Rational

choice may require comparability – explicit all-things-considered comparisons of the alternatives (but see Andreou, 2020). But rational choice does not require commensurability, and certainly not monetary commensurability. In response to these logical claims, many economists retreat from insisting that rational public decision-making necessitates monetary commensurability, and instead defend it on pragmatic grounds. In essence, the pragmatic defence is that monetary valuation of all environmental impacts leads to better environmental policy making. Clearly this defence raises a wide range of issues that are beyond our scope here. However, it is worth noting that many of the arguments in support of ‘incommensurable values’ in an environmental context are ultimately pragmatic arguments that monetary valuation of all environmental impacts does not provide useful information for policy making. ‘Incommensurable’ here does not mean that monetary valuation is impossible; rather, the claim is that since the monetary value of some environmental impacts is not meaningfully defined, or at best is extremely hard to measure, then monetary valuation is not informative. Objections to monetary valuation that focus on the absence of coherent preferences, or doubt the relevance or reliability of answers to contingent valuation survey questions, are examples of pragmatic arguments that monetary valuation is uninformative. A more direct response to the argument for monetary valuation on pragmatic grounds is to point to policy making that seems to function effectively without monetary valuation techniques. A cursory study of policy making practice in most countries reveals a wide range of policy arenas in which CBA, and related methodologies using monetary valuation, are almost never used. In health economics, a mainstream view is to evaluate the benefits to patients of various treatments in terms of Quality Adjusted Life Years (QALYs) rather than in monetary units. In many public policy arenas as diverse as foreign policy and abortion law, it is widely accepted that not all impacts can be evaluated in monetary terms. The pros and cons of different policy options are systematically evaluated, but these evaluations are presented in their natural quantitative units where possible, and in qualitative terms otherwise. In ecoJonathan Aldred

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logical economics, there is growing support for deliberative and participatory methods as an alternative to eliciting monetary valuations expressed in surveys or markets. These possibilities raise two related questions. First, what is the proper scope of monetary valuation? Second, what is the appropriate framework for answering this question? For example, most monetary valuation techniques are attempting to elicit preferences of consumers in markets or surrogate markets, whereas deliberative and participatory methods are attempting to elicit judgements from citizens in a political forum. In conclusion, in terms of policy making in practice, it is hard to separate debates over incommensurable values from these two wider questions. These questions have received remarkably little discussion among economists. Debates over incommensurable values raise a number of complex issues in ethics and economics. Ecological economists defending ‘incommensurable values’ in an environmental policy context might do better to focus directly on questioning the scope of monetary valuation. Why must environmental policy use monetary valuation when many other policy arenas do not? Supporters of monetary valuation have not yet answered this question. Jonathan Aldred

References

Andreou, C., 2020. Empowering rationality. American Philosophical Quarterly 57, 105–16.

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Brekke, K., 1997. The numeraire matters in cost– benefit analysis. Journal of Public Economics 64, 117–23. Claro, E., 2007. Exchange relationships and the environment: the acceptability of compensation in the siting of waste disposal facilities. Environmental Values 16, 187–208. Frank, R., 2001. Why is cost–benefit analysis so controversial?, in: Adler, M., Posner, E. (Eds.), Cost–Benefit Analysis: Legal, Economic and Philosophical Perspectives, 77–94. Chicago University Press. Frey, B., Jegen, R., 2001. Motivation crowding theory. Journal of Economic Surveys 15, 589–611. Frey, B., Stutzer, A., 2008. Environmental morale and motivation, in: Lewis, A. (Ed.), The Cambridge Handbook of Psychology and Economic Behaviour, 406–29. Cambridge University Press. Grant, R., 2012. Strings Attached. Princeton University Press. Hansson, S., 2007. Philosophical problems in cost– benefit analysis. Economics and Philosophy 23, 163–84. Hausman, D., McPherson, M., 2009. Preference satisfaction and welfare economics. Economics and Philosophy 25, 1–25. Layard, R., Glaister, S., 1994. Cost–Benefit Analysis. Cambridge University Press. Lukes, S., 1996. On trade-offs between values, in: Farina, F., Hahn, F., Vannucci, S. (Eds.), Ethics, Rationality and Economic Behaviour, 36–49. Clarendon. Mansfield, C., Van Houtven, G., Huber, J., 2002. Compensating for public harms: why public goods are preferred to money. Land Economics 78, 368–89. O’Neill, J., 2007. Markets, Deliberation and Environment. Routledge. Sen, A., 1997. Maximization and the act of choice. Econometrica 65, 745–79.

53. Industrial ecology As an inter- and transdisciplinary approach to understanding (un)sustainability, industrial ecology understands economic systems as embedded in their natural environments and, on this basis, has developed concepts and tools that are highly compatible and sometimes even overlapping with ecological economics. Much like ecological economics, industrial ecology is also a heterogeneous field of research and harbors a productive plurality of approaches to society–nature relations. Rather than providing a definition of a monolithic industrial ecology, this entry therefore seeks to delineate the research field and to provide examples for how its boundaries have shifted. The term “industrial ecology” initially appeared as a metaphor to capture intricate relations in newly emerging industrial systems. This metaphor was eventually adopted to describe a field of research in which industrial ecology is 1) the study of the “ecology” of industrial systems, that is, of their constituents and the relationships between them; and/or 2) the study of ecosystems and the environment from the perspective of industrial production. These loosely defined and overlapping research clusters both generate knowledge for socio-ecological transformations.

The ecology of industrial systems

All industrial systems are embedded in the environment: through the resources they require, the material stocks they accumulate, the wastes and emissions they generate. In a biological analogy, the constituents of an industrial system can be understood to have, like the constituents of an ecosystem, a social metabolism on which their reproduction depends: a need for material and energy inputs, transformation, stocks, and outputs (Fischer-Kowalski, 1998; Fischer-Kowalski and Hüttler, 1998). Industrial ecology has developed environmental accounting tools through which this metabolism can be tracked using standardized approaches and thereby generating comparable results. Because these tools can be operationalized at different levels of scale, they – material and energy flow accounting, substance or material flow analysis, life cycle analysis, and environmen-

tally extended input–output analysis – have many practical applications, from assessing the resource efficiency or circularity of entire economies to evaluating the environmental and even social implications of specific production or consumption processes. For a given socio-economic system, such as a city, a national economy, or a global region, material and energy flow accounting (MEFA) tracks inputs, transformation and accumulation, and outputs. The system obtains inputs either from its domestic environment through extraction (harvest, mining, hunting and gathering) or from other socio-economic systems (imports). These flows may then be accumulated in societal stocks (buildings and infrastructures, machines, durable consumer goods, humans and their livestock) or be rapidly used and transformed into wastes and emissions, as one form of output. The second form of output, to other socio-economic systems, is exports (Fischer-Kowalski et al., 2011). MEFA is an economy-wide approach in that it tracks all flows crossing the system boundaries; this makes it particularly useful in assessing not only the level and composition of societal resource use, but also in tracking trade-offs and synergies across different types of flows. Substance flow analysis (SFA, sometimes also referred to as material flow analysis) is also an operationalization of the concept of social metabolism and traces a specific chemical compound (or group of compounds) through a defined system – considering, as MEFA does, its inputs, stocks, and outputs. SFA is an adaptation of the Earth system sciences’ analyses of biogeochemical cycles at the level of economies, industries, or production processes. Through the focus on substances, SFA allows for the detailed assessment of environmental impacts as well as of possibilities for and limits to the use of less harmful substitutes (van der Voet, 2002). This type of analysis also makes it possible to trace flows not only into and out of the industrial system but also within that system, allowing for the identification of relevant actors and synergies or trade-offs with other processes governing the substance flows (e.g., Matsubae-Yokoyama et al., 2009). An equally detailed analysis, although from a different perspective, is provided by life cycle assessment (LCA). LCA considers all resource flows – from resource extraction to wastes and emissions – associated with

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the entire life cycle of a given product or service. This tool has been and is developed and applied in a transdisciplinary context involving businesses or governments seeking to understand the environmental impacts associated with certain production processes as well as consumers assessing the impact of their consumption choices (Curran, 2012; Hellweg and Canals, 2014). With an eye on providing knowledge for a transformation toward not only environmental sustainability but also greater social justice, the tool of social LCA has been introduced, which considers the social impacts along the life cycle of a product or service (Norris, 2012). Intensified globalization and increasingly fragmented international supply and use chains have given rise to the need to conceptually and empirically link final demand and consumption to resource inputs, wastes, and emissions, no matter where in the world they occur. This can be accomplished through footprint-type indicators (e.g., carbon footprint, material footprint, land footprint) calculated through environmentally extended input–output analysis (Duchin, 1992). This methodological innovation has allowed industrial ecology to make important contributions to understanding the role of subglobal patterns of production and consumption in determining global resource use (e.g., Plank et al., 2018) or the global implications of consumption at the household level (West et al., 2016). Considering the origin of inputs beyond the boundaries of a given socio-economic system is especially relevant to the study of cities and their urban metabolism (Weisz and Steinberger, 2010). Much of the greenhouse gas associated with the final demand in urban areas may, for example, occur outside the city boundaries and can only be accounted for as part of the urban metabolism through footprint-type approaches (Ramaswami et al., 2012).

The industry of ecological systems

If industrial systems can be better understood and eventually transformed from an ecological perspective, then studying how nature accomplishes processes that are also relevant to industry (flow and stock balance and sustainability at the system level, material cycling, mutually beneficial relationships between constituents) can inform the transformation of said systems. In adopting this Anke Schaffartzik

perspective, industrial ecology has given rise to a wealth of research on the characteristics of ecosystems and “designs” found in nature that are conducive to sustainability and resource-sparing, and on how such observations might be translated into greater sustainability of industrial processes. Such research spans the notion of biomimicry and the concept of industrial symbiosis, which can be understood as forming the foundation for the strategies for circular economies and/ or bioeconomies. The idea of biomimicry acknowledges that there are, within nature, solutions to problems of industrial systems. For example, plant leaves provide a design blueprint for optimal sunlight capture, and fungal growth represents optimal connections between given nodes. While the former already plays an important role in the design of solar panels, the latter can inform transport system design. In this sense, the biomimicry perspective investigates how naturally occurring forms might inform industrial design and process planning (Benyus, 2009). If this perspective is taken to ecosystems as a whole, then it becomes apparent that, compared to industrial systems, they are characterized by less waste and pollution and by greater sustainability over time, based on continuous, dynamic reconfigurations. In an analogy to biological symbiosis, industrial ecology has developed the concept of industrial symbiosis to capture how traditionally separate or disparate production processes might benefit from the mutual exchange of resource and waste flows (Chertow, 2000), and industrial systems have been interpreted as food webs (Hardy and Graedel, 2002). In nature, one species’ waste (e.g., dead organic matter, excrements) is generally another’s resource (e.g., food for animals, fungi, bacteria), and linear processes rarely exist. For biomass-based industrial production systems, the idea of cascade utilization of agricultural harvest for food, feed, fuel, and even fertilizer (Liu et al., 2019) can be considered symbiotic. Structurally, this type of exchange can be realized in so-called eco-industrial parks in which symbiotic processes can occur in close spatial proximity to one another. The currently emerging strategies for a circular economy (CE) are based on industrial ecology’s observations of natural cycling (rather than linearity; Saavedra et al., 2018; Winans et al., 2017). In a CE, ideally, the

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wastes and emissions of one industrial process become inputs for another, thus avoiding waste and emissions and decreasing the need for virgin resources. While ecological economics research has pointed to limits to the potential circularity of many processes which drastically increase entropy and would require substantial energy inputs to decrease this measure of thermodynamic disorder, industrial ecology’s research on symbiosis points to areas in which greater circularity may be feasible.

Industrial ecology for socio-ecological transformations

The key tools and concepts of industrial ecology certainly do not come with a prescription regarding the ends to which they may be employed. While they all provide opportunities to demonstrate where and how resource flows might be reduced and reallocated, it is also entirely feasible to use such approaches to improve efficiency and sustain economic growth, in the vein of environmental economics’ optimization. Much of industrial ecology research, however, is explicitly dedicated to generating knowledge for socio-ecological transformations. This endeavor, it has been noted, requires pushing the boundaries of the field, especially in recognizing the roles of actors and institutions involved and the power relations between them. Eco-industrial parks do not require only the technical possibility of symbiotic relationships across industrial processes in order to function. Instead, they necessitate the cooperation and a certain level of trust between actors who might simultaneously find themselves in (economic) competition with one another: a constellation that clearly requires stronger research engagement with actors and their relations on the part of industrial ecology (Ashton, 2008). In ensuring contributions to knowledge for a transformation that is not “only” environmentally sustainable but also socially just, industrial ecology is also called upon to explicitly recognize the role of power relations in governing and otherwise determining current patterns and levels of resource use (Breetz, 2017). This focus has been instrumental in combining the tracking of resource

flows with an analysis of the actors and institutions who have decision-making power over said flows (e.g., Deutz et al., 2017). Continuous innovations within the field of industrial ecology are what ensure that the wealth of analytical tools and practical concepts are employed toward advancing a socio-ecological transformation for a socially just and environmentally sustainable society. Anke Schaffartzik

References

Ashton, W., 2008. Understanding the organization of industrial ecosystems. Journal of Industrial Ecology 12, 34–51. https://​doi​.org/​10​.1111/​j​ .1530​-9290​.2008​.00002​.x Benyus, J.M., 2009. Biomimicry: Innovation Inspired by Nature. Harper Collins. Breetz, H.L., 2017. Political-industrial ecology: integrative, complementary, and critical approaches. Geoforum 85, 392–5. https://​ doi​ .org/​10​.1016/​j​.geoforum​.2016​.11​.011 Chertow, M.R., 2000. Industrial symbiosis: literature and taxonomy. Annual Review of Environment and Resources 25, 313–37. https://​ doi​.org/​10​.1146/​annurev​.energy​.25​.1​.313 Curran, M.A., 2012. Life Cycle Assessment Handbook: A Guide for Environmentally Sustainable Products. John Wiley & Sons. Deutz, P., Baxter, H., Gibbs, D., Mayes, W.M., Gomes, H.I., 2017. Resource recovery and remediation of highly alkaline residues: a political-industrial ecology approach to building a circular economy. Geoforum 85, 336–44. https://​doi​.org/​10​.1016/​j​.geoforum​.2017​.03​ .021 Duchin, F., 1992. Industrial input–output analysis: implications for industrial ecology. Proceedings of the National Academy of Sciences (PNAS) 89, 851–5. https://​doi​.org/​10​.1073/​pnas​.89​.3​ .851 Fischer-Kowalski, M., 1998. Society’s metabolism: the intellectual history of material flow analysis, Part I: 1860–1970. Journal of Industrial Ecology 2, 61–78. https://​doi​.org/​10​ .1162/​jiec​.1998​.2​.1​.61 Fischer-Kowalski, M., Hüttler, W., 1998. Society’s metabolism: the intellectual history of material flow analysis, Part II: 1970–1998. Journal of Industrial Ecology. 2, 107–36. https://​doi​.org/​ 10​.1162/​jiec​.1998​.2​.4​.107 Fischer-Kowalski, M., Krausmann, F., Giljum, S., Lutter, S., Mayer, A., Bringezu, S., Moriguchi, Y., Schütz, H., Schandl, H., Weisz, H., 2011. Methodology and indicators of economy-wide material flow accounting. Journal of Industrial

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312  Elgar encyclopedia of ecological economics Ecology 15, 855–76. https://​doi​.org/​10​.1111/​j​ .1530​-9290​.2011​.00366​.x Hardy, C., Graedel, T.E., 2002. Industrial ecosystems as food webs. Journal of Industrial Ecology 6, 29–38. https://​doi​.org/​10​.1162/​ 108819802320971623 Hellweg, S., Canals, L.M., 2014. Emerging approaches, challenges and opportunities in life cycle assessment. Science 344(6188), 1109–13. https://​doi​.org/​10​.1126/​science​.1248361 Liu, Y., Nie, Y., Lu, X., Zhang, X., He, H., Pan, F., Zhou, L., Liu, X., Ji, X., Zhang, S., 2019. Cascade utilization of lignocellulosic biomass to high-value products. Green Chemistry 21, 3499–3535. https://​doi​.org/​10​ .1039/​C9GC00473D Matsubae-Yokoyama, K., Kubo, H., Nakajima, K., Nagasaka, T., 2009. A material flow analysis of phosphorus in Japan. Journal of Industrial Ecology 13, 687–705. https://​doi​.org/​10​.1111/​j​ .1530​-9290​.2009​.00162​.x Norris, C.B., 2012. Social life cycle assessment: a technique providing a new wealth of information to inform sustainability-related decision making, in: Life Cycle Assessment Handbook, M.A. Curran (ed.). John Wiley & Sons, pp.  433–51. https://​doi​.org/​10​.1002/​ 9781118528372​.ch20 Plank, B., Eisenmenger, N., Schaffartzik, A., Wiedenhofer, D., 2018. International trade drives global resource use: a structural decomposition analysis of raw material consumption from 1990–2010. Environmental Science &

Anke Schaffartzik

Technology 52(7), 4190–98. https://​doi​.org/​10​ .1021/​acs​.est​.7b06133 Ramaswami, A., Chavez, A., Chertow, M., 2012. Carbon footprinting of cities and implications for analysis of urban material and energy flows. Journal of Industrial Ecology 16, 783–5. https://​ doi​.org/​10​.1111/​j​.1530​-9290​.2012​.00569​.x Saavedra, Y.M.B., Iritani, D.R., Pavan, A.L.R., Ometto, A.R., 2018. Theoretical contribution of industrial ecology to circular economy. Journal of Cleaner Production 170, 1514–22. https://​doi​ .org/​10​.1016/​j​.jclepro​.2017​.09​.260 van der Voet, E., 2002. Substance flow analysis methodology, in: A Handbook of Industrial Ecology, R.U. Ayres and L.W. Ayres (eds.). Edward Elgar Publishing, pp. 91–101. Weisz, H., Steinberger, J.K., 2010. Reducing energy and material flows in cities. Current Opinion in Environmental Sustainability 2, 185–92. https://​doi​.org/​10​.1016/​j​.cosust​.2010​ .05​.010 West, S.E., Owen, A., Axelsson, K., West, C.D., 2016. Evaluating the use of a carbon footprint calculator: communicating impacts of consumption at household level and exploring mitigation options. Journal of Industrial Ecology 20, 396–409. https://​doi​.org/​10​.1111/​ jiec​.12372 Winans, K., Kendall, A., Deng, H., 2017. The history and current applications of the circular economy concept. Renewable and Sustainable Energy Reviews 68, 825–33. https://​doi​.org/​10​ .1016/​j​.rser​.2016​.09​.123

54. Institutions 54.1 Introduction

54.2

Institutions play a key role in structuring relations between actors. They define what is normal or expected in certain situations of life. This way, institutions form individual and collective actors, such as organizations and their actions. In everyday language, the concept of an institution often refers to elderly homes, universi­ties, or public offices. This entry is based on a different understanding. It follows the tradition in, for example, institutional economics and sociology that defines institutions as ‘rules’ or ‘ways of doing things’ that are common to people. This maintains a distinction between institutions and actors where the latter also includes organizations. Institutions play an important role with regard to the status of our bio-physical environments. They influence who gets access to various environmental resources. Next, they affect the use of these assets, how and what we produce and consume. The combination of profit-seeking firms operating in (global) markets for commodities has resulted in a tremendous increase in the use of natural resources. While this has facilitated, enhanced, but also very unevenly distributed material well-being, it has resulted in tremendous pressures on our bio-physical environments. These environments are common to us, and the actions of one cannot be isolated from the actions of others. Nevertheless, to facilitate economic growth, we have developed institutional struc­tures support­ing independent action through the strengthening of private ownership and trade of commodi­ties in liberalized markets. The independencies thus created are, however, only nominal. The independent choices facilitated via commodity trade are still connected via the environment – both through the required use of natural resources and the necessary creation of waste. Through facilitating economic growth, our interdependencies are greatly elevated. As the problem is institutional, its solution also lies in changing our institutions.

The concept of an institution

So, how do we define an institution? Despite limiting the concept to that of ‘common rules’, the number of definitions found in the literature is vast. A few are included below, starting with that of Thorstein Veblen – the ‘father’ of institutional economics (emphasis added for each): ● Veblen (1919: 239): (Institutions are) settled habits of thought common to the generality of man. ● North (1990: 3): Institutions are the rules of the game in a society or, more formally, the humanly devised constraints that shape human interaction. ● Scott (2014: 56): Institutions comprise regulative, normative, and cultural-cognitive elements that, together with associated activities and resources, provide stability and meaning to social life. Veblen developed his position much as a reaction to neoclassical economics. North was a leading scholar within the so-called new institutional economics tradition, while Scott is a sociologist working within the field of organization studies. His position overlaps, to a considerable extent, the tradition originating in Veblen. The literature accentuates that institutions are human constructs – here made explicit in the definition by North. Otherwise, there is a crucial difference illustrated by the chosen examples. While North emphasizes that institutions are external constraints, both Veblen and Scott see institutions as also forming by people. This reflects a different understanding of the human. While North is an institutional economist, he still retains the neoclassical understanding that people’s preferences are specific to the individual. Hence, institutions can only take the form of constraints. The alternative view held by authors like Veblen and Scott is that insti­tutions become internalized through processes of socializing. They shape who we become. This distinction is formative of the two main traditions in institutional economics – that of classic and new institutional economics. The former ‘school’ was initiated by Veblen more than a hundred years ago. Other more recent representatives encompass scholars like Hodgson (1988), Bromley

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(1989), and Schmid (2004). The latter tradition has a shorter history, and one may track its commence­ ment to Coase (1960), while important developments are found in Williamson (1985) and North (1990). Language is the most important institution as it forms the basis for all other institutions. It is based on a set of rules – syntax and semantics – that makes it possible to transfer meaning­ful messages. Parallel to Scott’s distinction between cultural-cognitive, nortive elements, one may mative, and regula­ further distinguish between three forms of institutions: conventions, norms, and formal rules (see also Crawford and Ostrom, 1995; Vatn, 2005). Conventions are rules about how things normally are, or are normally done. They create order in society and simplify interaction. A wide variety of examples beyond language include, among others, measure­ ment scales, how to perform a certain task or profession, dress codes, and on which side of the road to drive. In this sense, conventions are typically formed when issues are non-conflictual. The issue is to have a rule, not so much what it is (e.g. whether we drive on the right or left side of the road). Conventions also encompass ‘ways of doing things’. They are ‘condensed’ experiences, like those forming a profession. Creating conventions, moreover, ‘natur­al­izes’ human tions. They form our cognition of the rela­ external world (Scott, 2014) and define what is ‘normal’. This takes us to another aspect of conventions. Orders are often not neutral, and conflicts may be made invisible by ‘naturalizing’ uneven power relations and hierarchical structures (cf. Lukes’, 2005, third dimension of power). The Indian caste system is a prime exam­ple. People are shudras or brahmins, and even the former believe that they are of a lower status than the latter. Also, power structures under capitalism incorporate many such elements. Norms are distinguished from conventions as they refer to a required act or solution – something one should or should not do. In that sense, norms protect an underlying value. Examples would be the norm of greeting others,1 general norms regarding care, and norms against littering and lying. Again, the list could be made very long. The main point is that norms are about how to treat others – both other humans and other-than-humans. Norms take sides and protect certain interests Arild Vatn

or values. Hence, there is an element of (potential) conflict involved. Formal or legal rules also regard something we should or should not do. Different to a norm, a third-party power is instituted, having the authority to both formulate and enforce this kind of institution. Such power could be vested in a state or in more traditional authorities. Formal rules are typically instituted when there is potential for high levels of conflict. Examples are property and use rights, rights to tax, regulations of activities creating environ­mental deterior­ation, and responsibilities for public service delivery. The distribution of property rights is crucial for the opportunities people face and has strong impacts on the distribution of income – hence, also power – in societies. Going from conventions to formal rules, the power to influence becomes more explicit. As we have seen, that may not necessarily imply that conventions are devoid of power relations. Also, norms are expressions of power in the sense that they favour one value above another. However, it is the general acceptance in a culture that makes it remain a norm and not (trans-)formed into a legal rule. The difference between the two lies in the type of authority structure involved. The field of institutional theory is complex, not least because authors define and conceptualize institutions differently. We have categorized institutions as conventions, norms, and legal rules. Veblen defined institutions as habits – as automatic and durable actions. Moreover, they become ‘who we are’. Habituation is therefore a process of enculturation and identity formation (i.e. Veblen’s formulation of the ‘generality of man’). Classical institu­ tional economists share this insight (see also Hodgson, 2010). While the naming of what an institution is may differ, the perspective on how they form us is common. Before closing the discussion regarding definitions, we should visit social psychology. The concept of a norm is important in that tradition (e.g. Cialdini et al., 1991). Here the main distinction is between social and personal norms. Social norms refer to expectations by others about what one should/should not do. Personal norms, on the other hand, are the norms held by an individual based on the individual’s understanding of what they deem important. Cialdini et al. (1991) differentiate social norms as ‘injunctive’ and ‘descrip­tive’

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social norms. The former refers to explicit expectations by others about what one should or should not do. The latter refers to what others do. It is worth noting these differences, as many authors in the social sciences use the concept of an institution seemingly without paying attention to the different epistemological and ontological bases involved. A social norm (in psychology) is partly equivalent to a non-internalized norm (the injunctive part) and partly to a convention (the descriptive part) in institutional econ­ omics/sociology. Both tradi­tions acknowledge that social processes are important, but social psychologists are less engaged in the role institutions play in forming individuals. That is most clear if we look at the concept of ‘a personal norm’. As noted, social psychologists see these as individual characteristics, as chosen by the individual. Classical institutional economists and sociologists, on the other hand, define these as internalized norms – as examples of encultura­tion/how society imprints itself upon the individual. So, classical institutional economics holds that institutions both form and are formed by humans. In the case of formal rules, formation happens, foremost, as a result of political pro­ cesses (e.g. decisions in parliaments). Here, power dynamics, interests, and value conflicts often become visible, and the influence of different actors on, for example, a new law is traceable to some extent. Regarding conventions and norms, change is much more ‘dispersed’. Some norms may develop as a result of public debates, and one can somehow understand what happened – what was the reason(s) for a norm change. Many conventions, such as consumption patterns, are harder to trace. Certainly, people holding visible positions in society may influence these patterns due to our tendency to copy others. In modern society, business also plays a key role here. Still, understanding why a specific convention ‘won’ may be hard to disen­tangle. The fact that people may also move between societies enhances the complexity of institutional change. Some may be unhappy with the ‘rules’ prominent in one society and move to or form another. This fact may be seen to support the social psychological perspective on personal norms – as being about individual choice. A classical insti-

tutionalist would maintain that there is still a very important element of enculturation and collective choice involved. Moving to another society implies becoming part of that society with its conventions and norms. Creating a new ‘culture’ is, moreover, a collective endeavour. It happens through com­ munication and agreement with others.

54.3

Institutions and human action

The key aspect of institutions is to influence our actions. That is true whether they are seen only as external constraints or as (also) forming and enabling us. In the former case, institutions work through a calculative form of logic: Is it worth following the norm or legal rule? Classical institutional economists do not deny that an effect of institutions is to constrain people. They, however, also emphasize their liberating role, such as expanding our capacities and opportuni­ties (e.g. Bromley, 2006). In this literature we also observe an emphasis on how institutions influ­ ence the very logic of action (March and Olsen, 1995) and what is found rational to do (Vatn, 2005; Hodgson, 2007; Sandel, 2013). Even the results of experiments in behavioural economics – a tradition coming out of neoclassical economics – may be taken to support such an understanding. Hodgson (2007: 329) concludes that this research has ‘revealed the limi­tations of all-purpose, context-independent rationality and pointed to the institutional influ­ences on rationality itself’. One way to understand this is to see institutions as rationality contexts. In certain settings, institutions emphasize what is individually best to do – like in markets – while in other situa­tions they emphasize what is better for the community of people (e.g. Etzioni, 1988; Vatn, 2005). This is very explicit in the forming of the role of a representative, like a politician or an administrator, whose responsibility is to act on behalf of a community or nation (March and Olsen, 1995). Hence, we can both talk about individual and social rationality. The former regards what is best for me (I rationality) while the latter regards what is better for the group one belongs to (we rationality/solidarity) or for others (they rationality/altruism). Arild Vatn

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According to this perspective, societies define different arenas for action. In some cases, these protect the right of individuals and organizations to do what they want. In other situations, one’s (potential) impacts on the opportunities of others are deemed unacceptable, and norms or legal rules are created emphasizing what is considered better for the group or for others. The literature is full of examples documenting how shifts in the institutional structures change the logic from ‘what is in it for me’ to ‘what is appropriate’, or the other way around (e.g. March and Olsen, 1995; Gneezy and Rustichini, 2000; Bowles, 2008; Kerr et al., 2012). Sandel (2013: 128) points out that markets influence our actions both by promoting a certain form of valuing goods and ‘erod[ing] or crowd[ing] out nonmarket motivations’. Two caveats should be added to the above. First, even what is individually rational – our capacities to act rationally as well as what we want – are socially influenced (Veblen, 1899; Etzioni, 1988; Grusec and Hastings, 2007). Hence, societies do not only influalities pertain to specific ence what ration­ situations. They also influence what these rationalities mean in more concrete terms. As an example, they do not only say ‘you should not litter’, but they define what avoidance of littering means (e.g. conventions for sorting waste). Second, there is not only institutional variation. There are also differences across individuals as people may follow different con­venti­ons or abide by norms to a different degree. This may reflect different upbringing as well as genetic variations.

54.4

Institutions, governance, and the environment

As we have seen, institutions play a fundamental role in shaping how we act and the dynamics of action. They also influence heavily our access to environmental resources and the distribution of this access. These are important insights for ecological economists, given the strong emphasis on limiting our collective impacts on the bio-physical basis for our lives. The Earth system (Rockström et al., 2009) is a system based on interdependent processes, with subsystems often characterized by tipping points. As I have emphasized above, our actions have, however, been made institutionally more independent and, over time, we experience that the sum of Arild Vatn

such (nominally) independent actions may be devastating (Rockström et al., 2009; Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services [IPBES], 2019; Gómez-Baggethun, 2020). Therefore, a key issue for environmental policies and governance regards how to create societal systems where it is possible to limit the collective consequences of various actions. Institutional theory points towards redistribution of access to resources and strengthening the role of social rationality. This implies shifting the direction of institutional developments from reinforcing the role of individual ownership and choice (e.g. neoliberal market expansions) to facilitating the role of social rationality at all levels of society. This is very demanding. Briefly disentangling the challenges, one may start by noting that governance is about engaging not only government but also civil society and business (Evans, 2012; Vatn, 2015). Hence, in governance we can distinguish among political actors, economic actors, and civil society actors, all with their specific institutions defining what is considered ‘normal’ or acceptable actions. Political actors – such as parliaments, governments, and municipal boards – have the power to define and protect rights to environmental resources and how these can be used. These decisions follow responsibilities as defined in the constitution and in collective choice rules (Ostrom, 1990). Economic actors use resources given the defined institutional structures of rights and regulations regarding these resources. They follow what may be called operational rules, defining how the actual resources can be used and possibly traded (Ostrom, 1990). While corporations are dominated by the goal of profit maximization (e.g. Sjåfjell, 2011), community and family-based production is observed to follow a wider set of aims emphasizing, for example, community values and reciprocity (Bowles, 1998). Finally, civil society elements (as e.g. manifested in nongovernmental organizations, political parties, mass movements) engage citizens in discussions about what are reasonable solutions to issues faced (Dryzek, 2010). It is at this level that the legitimacy of political action is formed (e.g. Habermas, 1991). An important step regarding needed institutional change could be to strengthen civil society engagement. That could be fostered

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by establishing deliberative, participatory forums across society, such as citizens’ assemblies (e.g. Dryzek, 2010; Wironen et al., 2019). Following the perspectives of March and Olsen (1995), formulating mandates for these assemblies directed at creating ideas about how to establish sustainable trajectories of the economy and society more at large would influence their focus greatly. Political decision-making is criticized for being short-sighted (González-Ricoy and Grossieres, 2016) and emphasizing narrow political goals like economic growth causing grave environ­mental destruction (Hickel and Kallis, 2019; Spash, 2020). Again, institutional change could help strengthen the role of the long-term and wider conceptualizations of welfare rather than maximizing monetary income. This could happen through constitutional changes strengthening equality and the rights of the future. Maybe more importantly, it could be strengthened by changing procedures of political decision-making – the collective choice rules – where long-term sustainability is given prominence when forming institutions for the economic sphere. The idea of a ‘sustainability chamber’ in the parliament illustrates the point (Vatn, 2020). Finally, it will be important to increase the role of sustainability in the business sector. While businesses are largely focused on earning the highest possible profits and are forced to compete in markets – often implying negative consequences for the biosphere – there are also examples of business forms that build their mission on goals and values where protecting nature is important (e.g. so-called eco-social enterprises; Johanisova and Franková, 2017). Increasing the importance of such forms of production demands both legal and financial support. Moreover, these developments illustrate that social rationality is a possible logic also for business activities. Strengthening the position of community ownership and cooperatives may expand such a development, also ensuring more equal access to resources. The above represents brief illustrations of the role institutions and institutional change could play in ensuring sustainable futures. Institutional change is important since institutions influence people, the relations between people, and between people and the bio-physical environ­ments that societies depend on. While the most basic question

regards what values we want to foster and protect, the subsequent question regards what institutions that make these values materialize could look like. Arild Vatn

Note 1.

Although the way greeting is done is better categorized as a convention.

References

Bowles, S., 1998. Endogenous Preferences: The Cultural Consequences of Markets and Other Economic Institutions. Journal of Economic Litera­ture, XXXVI (March):75–111. Bowles, S., 2008. Policies Designed for Self-Interested Citizens May Undermine “The Moral Sentiments”: Evidence from Economic Experiments. Science, 320(5883):1605–09. Bromley, D.W., 1989. Economic Interests and Institutions. The Conceptual Foundations of Public Policy. Oxford: Basil Blackwell. Bromley, D.W., 2006. Sufficient Reason: Volitional Pragmatism and the Meaning of Economic Institutions. Princeton, NJ: Princeton University Press. Cialdini, R., Kallgren, C., and Reno, R., 1991. A Focus Theory of Normative Conduct: A Theoretical Refinement and Re-Evaluation of the Role of Norms in Human Behaviour. Advances in Experimental Social Psychology, 24:201–34. Coase, R.H., 1960. The Problem of Social Cost. Journal of Law and Economics, 3:1–44. Crawford, S.E.S., and Ostrom, E., 1995. The Grammar of Institutions. American Political Science Review, 89(3):582–600. Dryzek, J.S., 2010. Foundations and Frontiers of Deliberative Governance. Oxford: Oxford University Press. Etzioni, A., 1988. The Moral Dimension: Toward a new Economics. New York: The Free Press. Evans, J.P., 2012. Environmental Governance. London: Routledge Gneezy, U., and Rustichini, A., 2000. Pay Enough or Don’t Pay at All. Journal of Economic Behavior and Organization, 39:341–69. Gómez-Baggethun, E., 2020. More is More: Scaling Political Ecology Within Limits to Growth. Political Geography, 76:102095 González-Ricoy, I., and Gosseries, A. (eds.), 2016. Institutions for Future Generations. Oxford: Oxford University Press. Grusec, J.E., and Hastings, P.D. (eds.), 2007. Handbook of Socialization: Theory and Research. New York: Guilford Press. Habermas, J., 1991. The Structural Transformation of the Public Sphere: An Inquiry into a Category

Arild Vatn

318  Elgar encyclopedia of ecological economics of Bourgeois Society. Cambridge, MA: MIT Press. Hickel, J., and Kallis, G., 2019. Is Green Growth Possible? New Political Economy, 25(4):1–18. Hodgson, G.M., 1988. Economics and Institutions: A Manifesto for a Modern Institutional Economics. Cambridge: Polity Press. Hodgson, G.M., 2007. The Revival of Veblenian Institutional Economics. Journal of Economic Issues, XLI(2):325–40. Hodgson, G.M., 2010. Choice, Habit and Evolution. Journal of Evolutionary Economics, 20:1–18. Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services (IPBES), 2019. Global Assessment Report on Biodiversity and Ecosystem Services. https://­​ ­ipbes​.net/​global​-assessment Johanisova, A., and Franková, E., 2017. Eco-Social Enterprises. In: Spash, C.L. (ed.), Handbook of Ecological Economics. Routledge, Oxon, pp. 507–16. Kerr, J., Vardhan, M., and Jindal, R., 2012. Prosocial Behavior and Incentives: Evidence from Field Experiments in Rural Mexico and Tanzania. Ecological Economics, 73:220–27. Lukes, S. (2005 [1974]). Power. A radical view. New York: Palgrave Macmillan. March, J.G., and Olsen, J.P., 1995. Democratic Governance. New York: Free Press. North, D.C., 1990. Institutions, Institutional Change and Economic Performance. Cambridge: Cambridge University Press. Ostrom, E., 1990. Governing the Commons: The Evolution for Collective Action. Cambridge: Cambridge University Press. Rockström, J., Steffen, W., Noone, K., et al., 2009. Planetary Boundaries: Exploring the Safe Operating Space for Humanity. Ecology and Society, 14(2):32. Sandel, M.J., 2013. Market Reasoning as Moral Reasoning: Why Economists Should Re-Engage

Arild Vatn

with Political Philosophy. Journal of Economic Perspective, 27(4):121–40. Schmid, A.A., 2004. Conflict and Cooperation. Institutional and Behavioral Economics. Oxford: Blackwell Publishing. Scott, W.R., 2014. Institutions and Organizations, 4th ed. Thousand Oaks, CA: Sage Publications. Sjåfjell, B., 2011. Why Law Matters: Corporate Social Irresponsibility and the Futility of Voluntary Climate Change Mitigation. European Company Law, 8(2–3):56–64. Spash, C.L., 2020. ‘The Economy’ as if People Mattered: Revisiting Critiques of Economic Growth in a Time of Crisis. Globalizations, 18(7):1087–1104. Vatn, A., 2005. Institutions and the Environment. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Vatn, A., 2015. Environmental Governance. Institutions, Policies and Actions. Cheltenham, UK, and Northampton, MA: Edward Elgar Publishing. Vatn, A., 2020. Institutions for Sustainability— Towards an Expanded Research Program for Ecological Economics. Ecological Economics, 168:106507. Veblen, T., 1899. The Theory of the Leisure Class: An Economic Study of Institutions. New York: Macmillan. Veblen, T., 1919. The Place of Science in Modern Civilisation and Other Essays. New York: Huebsch. Williamson, O.E., 1985. The Economic Institutions of Capitalism. New York: Free Press. Wironen, M.B., Bartlett, R.V., and Erickson, J.D., 2019. Deliberation and the Promise of a Deeply Democratic Sustainability Transition. Sustainability, 11:1023.

55. Joint production Joint production1 is one of the basic concepts of ecological economics. It means that several outputs of a single production activity necessarily emerge together. From the application of thermodynamics it follows that all production is joint production if all input and output streams are taken into account (i.e. joint products are intrinsic to production processes). In fact, joint products are often wastes or polluting substances. The analysis of joint production has a long history in economic analysis. Considerations of joint production give rise to philosophical concerns relating to responsibility and knowledge. The concept of joint production is easily comprehensible and relevant to many issues of environmental and sustainability policy.

Introduction – what is joint production?

In a nutshell, joint production means that several outputs of a single production activity necessarily2 emerge together.3 In the refining of crude oil, for example, gasoline, kerosene, light heating oil, and other mineral oil products are jointly produced. In the refining process, harmful sulphurous wastes and carbon dioxide emissions are also necessarily generated. The occurrence of joint production is a characteristic of a particular production system. It refers to the material and physical base of the underlying production activity. It means that, in this particular production system, it is not possible to produce the principal output without its joint outputs. Supposing that the production activity is pursued for (at least) one intended principal product, the presence of joint production itself does not say anything about whether the jointly emerging by-products are desirable (i.e. goods) or undesired (i.e. wastes). In the vast majority of instances, however, while one or several products may be desirable, other outputs are undesired and may even be harmful wastes. When taking into account the full range of input and output streams of a production activity, it can be shown from thermodynamics that basically all production is joint production. Hence, joint production is a ubiquitous phenomenon. However, this fact, which is based on the physical and material foundation

of production, also has a broader dimension: the concept of joint production captures the particular characteristic of human activity which is the structural cause of many environmental problems – namely, that it always has unintended side effects.

Thermodynamics and the ubiquity of joint production

The usefulness of thermodynamics derives from its applicability to all real production processes, which are the basis of economic activity. The laws of thermodynamics lead us to recognise that the human economy is an open subsystem embedded in the larger, but finite, system of the natural environment (Boulding 1966; Georgescu-Roegen 1971; Daly 1977; Ayres 1978; Faber et al. 1983; and many more). Using the notion of joint production allows us to incorporate this insight about economy–environment interactions into ecological economics. This can be seen from the following argument. From a thermodynamic point of view, energy and matter are the fundamental factors of production. Every process of production is, at root, a transformation of these factors. Hence, production processes are subject to the laws of thermodynamics, which, in an abbreviated form, can be stated as follows: ● First Law: Energy and matter can be neither created nor destroyed (i.e. in an isolated system, matter and energy are conserved). ● Second Law: In every real process of transformation, a positive amount of entropy is generated. One can describe the process of production as a transformation of a certain number of inputs into a certain number of outputs, each of which is characterised by its mass and its entropy. From the laws of thermodynamics it then follows that every process of production is joint production (i.e. it results necessarily in more than one output; Faber et al. 1998; Baumgärtner 2000: Chap. 4). In particular, production processes which generate low-entropy desired goods necessarily and unavoidably jointly produce high-entropy waste materials. We can represent this thermodynamic constraint on real production processes as in Figure 55.1.

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Source:

Baumgärtner et al. (2001).

Figure 55.1 Production processes generating low-entropy desired goods necessarily and unavoidably jointly produce high-entropy waste materials

For example, in the production of iron, one starts from iron ore. To produce the desired product, iron, which has lower specific entropy than iron ore, one has to reduce the raw material’s entropy. This is achieved by employing a low-entropy fuel (e.g. coal), which provides the energy necessary for this process. From a thermodynamic point of view, one may therefore consider production as a shifting of high entropy from the raw material to the waste product. At the same time, it becomes apparent that the inputs are also joint in the sense that high-entropy iron and low-entropy fuel are complementary (cf. Christensen 1989: 28–9). Hence, the fundamental idea of joint production applies both on the input and the output side. In that sense, the concept of joint production can capture the essential thermodynamic constraints on production processes as expressed by the First and Second laws, through an easy-to-use and easy-to-understand economic concept. This holds for production in both economic systems and ecosystems. Joint production, therefore, is also a fundamental notion in ecology, even though it is not often expressed as such in that discipline. Organisms and ecosystems, as open, self-organising systems, necessarily take in several inputs and generate several outputs, just as an economy does. Indeed, such natural systems are the earliest examples of joint production. The concept of joint production embraces four central issues in ecological economics:

irreversibility; limits to substitution; the ubiquity of waste; and the limits to growth. Irreversibility is explicitly included within the above thermodynamic formalisation of joint production, as it is necessarily the case that the production process generates entropy and is therefore irreversible. Limits to substitution are also included, as the requirement that high-entropy materials inputs must be converted into lower-entropy desired goods requires that the material inputs be accompanied by an irreducible minimum of low-entropy fuels. The ubiquity of waste can be easily derived from the thermodynamically founded joint production approach. It follows from the necessity of jointly producing high entropy, which very often is embodied in undesired material and, hence, constitutes waste (e.g. CO2, slag, etc.). The combination of the above three issues leads to the notion of limits to growth, further emphasising the power and generality of the joint production concept for ecological economic analysis.

Joint production and economics

Economic analysis of joint production The analysis of joint production has a long tradition in economics. For example, Adam Smith, John Stuart Mill, Karl Marx, Johann Heinrich von Thünen, William Stanley Jevons, Alfred Marshall, Arthur Cecil Pigou, Heinrich von Stackelberg, John von Neumann, and Piero Sraffa devoted considerable effort to the study of joint production (Kurz 1986; Baumgärtner 2000: Chapters

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5–8). Within the substantial body of literature both on theory and applications of joint production, two cases are distinguished: all joint products are desired goods, and at least one output is undesired while at least one other is desired. While the former is the case that has received the most treatment in the literature, the thermodynamic discussion leads us to conclude that it is the second case that is of particular interest in ecological economics. The theory of joint production has been extensively developed in business administration (e.g. Dyckhoff 1996). For example, a range of methods has been developed to solve the resulting problems concerning the planning and cost allocation of joint production (Oenning 1997). Further, the quantitative relations between inputs and outputs in joint production can be described with input–output graphs, and one can use linear or non-linear algebraic systems generalising Koopman’s (1951) activity analysis. Joint production is important for balancing and managing the flows of material and energy (Spengler 1999), such as in engineering and chemistry. There is a range of theoretical results about the economics of joint production (cf. Baumgärtner et al. 2006: Section 1.3): joint production of private and public goods may reduce the usual problem of under-provision of public goods in a decentralised economy (Cornes and Sandler 1984). Under joint production of goods and polluting residuals, and making the realistic assumption that the assimilative capacity of the natural environment for these pollutants is limited, a steady state growth path does not exist (Perrings 1994; O’Connor 1993). Literature on general equilibrium theory has implicitly dealt with joint production when investigating the most general assumptions under which certain results hold, for example, the existence and efficiency of general equilibrium (Arrow 1951; Arrow and Debreu 1954; Debreu 1951, 1959; McKenzie 1959). Pigou (1920) and Lindahl (1919) conceived mechanisms to internalise negative externalities, thereby re-establishing optimality of any general competitive equilibrium. In the case of negative externalities exhibiting the character of public bads, however, this mechanism can only be established under very restrictive and unrealistic assumptions.

In summary, while modern economic theory has produced many interesting results concerning the existence and efficiency of equilibrium under joint production, in the case which is most relevant from the ecological-economic point of view – joint production of bads causing public negative externalities – we are essentially left with a negative result. Joint production and external effects Having identified joint production as a structural cause of many environmental problems, it is useful to highlight the connection with another prominent economic concept for modelling environmental problems – the concept of externality. In welfare-based environmental economics, joint production is typically modelled in an implicit way, mediated by the dis-utility generated to a third party by, for example, the emission of a joint product. Following a standard definition of an externality, “[an] externality is present whenever the well-being of a consumer or the production possibilities of a firm are directly affected by the actions of another agent in the economy. ... When we say ‘directly’, we mean to exclude any effects that are mediated by prices” (Mas-Colell et al. 1995: 352). In the externality approach, the relationship between the agent causing the effect and the agent affected is conceptualised as an issue of welfare loss of the person affected by the external effect. Hence, without affected economic agents and their valuation of the effect, an externality does not exist. One could, however, recast this relationship starting from the cause of the effect. Very often one would observe that the starting point is an unintended joint product. Therefore, we observe that there exists a duality between an explanation based on the effect – that is, the externality approach – and an explanation starting from the cause of the effect – that is, the joint production approach. We also note that welfare effects will only be taken account of once they have been experienced. As some effects of current joint production will only show up in the future, external effects are ex post matters. On the other hand, the concept of joint production can alert one to the potential of environmental harm; that is, considering joint production ex ante creates a motive for actively explor-

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ing as yet unknown potential welfare effects (Baumgärtner 2000: 293–4).

Joint production and philosophy

Joint production stresses that economic activity generally produces two kinds of output: the intended principal product and unintended by-products. We would expect, and indeed observe, that producers will focus their attention and energies on the former, while the latter will be largely ignored, at least to the extent permitted by legal constraints and social mores. This inattention to the undesired products raises two issues of a philosophical nature, one relating to responsibility, that is ethical, and one relating to knowledge, that is epistemological. Ethics Turning first to ethics, the thermodynamically necessary by-products bring with them new issues of moral responsibility. This becomes obvious if we consider the hypothetical case of single production where no by-products are generated. In such an idealised world, assuming the existence of perfect markets and a fair social and legal order, the ethical problem for producers of a desired product is narrowly limited as long as they trade their products on the market and obey the legal order. In contrast, joint production implies that economic activity, in addition to the intended products, also results in unintended outputs, which often go unnoticed. This lack of knowledge and attention often results in a social and legal order that neglects joint products. However, these joint products may be harmful, for example, to other producers, consumers, or to the natural environment. As a consequence, both the producer and the wider society demanding the desired principal product now face complex ethical problems. Inattention to joint production may therefore easily result in ethical negligence. An example is the inattention to waste in the nuclear industry. From the inception of nuclear power, it was recognised that very dangerous and long-lived waste materials would be produced as by-products. Nevertheless, for the first 30 years of commercial power generation, unconscionably little attention was paid to the disposal of this waste (Proops 2001).

Epistemology Concerning the second issue, epistemology, the area to which we draw attention is that of surprise and ignorance (Faber et al. 1992). Even if one were to suppose that it were possible to produce only principal products, this could still give rise to unanticipated and unwanted environmental effects (for example, chlorofluorocarbons used to be a principal product, not a by-product). However, we believe that unwanted waste by-products are likely to be a greater source of unpleasant environmental surprises because, as mentioned above, they are not the focus of attention for their producers. The story of waste chlorine in the 19th century is one of ignorance of, and inattention to, the effects of emitting this waste product, with damaging and unforeseen consequences for air and water quality (Baumgärtner et al. 2006: Chapter 16). In summary, considering the concept of joint production naturally leads one to address issues of ethics and epistemology, requiring one to discuss economic questions in a philosophical context. In particular, the concept creates an awareness of both (i) the ethical dimension of economic action due to unintended joint outputs, and (ii) our potential ignorance, primarily of the effects of unwanted by-products.

Relevance of joint production for policy

Joint production draws our attention to the fact that economic activity is inextricably linked to side effects. Naturally, this leads to issues that are relevant for environmental and sustainability policy. The universality of the concept The concept of joint production may be employed at several different levels of aggregation. It can be used for the analysis of an individual production process, of a firm, of an economic sector, or of a whole economy. It is also suited to examine environment–economy interactions in which economic activities and resulting environmental effects are separated by long time intervals, as in the example of CO2 emissions. In both cases, today’s effects on the natural system are caused by stocks of these substances that were accumulated

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mainly from emissions up to several decades ago. Holistic approach to policy Taking a joint production approach to economy–environment interactions stresses the necessary relationships between various sorts of inputs into production processes, and the corresponding sorts of outputs. As illustrated in Figure 55.1, much, even most, production requires inputs of low-entropy fuels and high-entropy raw materials, and generates low-entropy desired goods and high-entropy wastes. Thus, this thermodynamically based joint production representation shows us that the two issues, of natural resource use and of pollution from waste, are necessarily and intimately related: the resource is the mother of the waste. So it is conceptually incomplete to consider natural resources and pollution as separate issues. The theory of joint production tells us that sound environmental policy can come only from an integrated and holistic conceptualisation of the production and consumption processes. Time scales and time horizons Joint production leads one to the recognition of different time scales and time horizons. Desired principal products are generally produced and consumed over relatively short time scales, leading to relatively short time horizons of decision-makers with regard to such outputs. However, jointly produced waste outputs are often emitted into the environment, where they can accumulate over longer time scales. Such accumulation may, and often does, lead to the unanticipated and unpleasant surprises. Clearly, the social management of such problems demands much longer time horizons than those typically applied to the principal products. Communication and awareness: joint production as a comprehensible principle It is clearly desirable that fundamental concepts of ecological economics should be easily comprehensible. It has often been noted in the literature (for example, by Norton 1992) that the scientific approach is sufficient neither for the recognition of environmental problems, nor for their solution. Concerning recognition, as a matter of history, the awareness of environmental deg-

radation was largely brought about not by the scientific community, but by laypeople. For, in everyday life, attentive human beings can recognise many dimensions of the natural environment, while science, by its nature, has to reduce the wholeness of an event to only those aspects to which its methods are suited. Concerning the solution, in democratic societies, decisions about what kind of environmental policy is to be enacted are made (effectively) by ballot. Hence, voters have to understand environmental issues and their proposed solutions. We have often noted in discussions with scientists who had no background in economics, but also with laypeople, that they were able to comprehend the nature of an environmental problem and to appreciate a proposed solution much more easily when such issues were explained in terms of joint production, rather than in other economic terms, for example, production functions, damage functions, externalities, Pigouvian taxes, and so on.

Conclusion

The concept of joint production can provide a translation of the insights from thermodynamics, which is notoriously difficult for those who are not trained in the field, into a language that can easily be understood. At the same time it allows ecologists to get in touch with economists and to make use of the large body of knowledge available in economics. In summary, joint production constitutes a foundational concept for ecological economics since: ● it is applicable to the natural systems with which humans interact, ● it is descriptive of economic activity, ● it relates to the areas of responsibility and human knowledge, and ● it is transparent and comprehensible to practitioners, policy makers, and the wider public. Hence, the concept of joint production unifies thermodynamic-ecological, economic, and philosophical principles. Viewing joint production in this way opens up directions for fruitful research drawing on various concepts and methods of economics and of the natural sciences. Johannes Schiller and Stefan Baumgärtner

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Notes 1. 2. 3.

This article, in significant parts, draws from material previously published in Baumgärtner et al. (2001 and 2006). The New Palgrave entry on joint production (Nadiri 1987) calls this necessity “intrinsic jointness”. A formal definition of joint production is given in Baumgärtner et al. (2006: Section 2.4).

References

Arrow, K.J., 1951. An extension of the basic theorems of classical welfare economics. In: J. Neyman (ed), Proceedings of the Second Berkeley Symposium on Mathematical Statistics and Probability. University of California Press, Berkely, pp. 507–32. Arrow, K.J., and G. Debreu, 1954. Existence of an equilibrium for a competitive economy. Econometrica, 22:265–90. Ayres, R.U., 1978. Resources, Environment, and Economics – Applications of the Materials/ Energy Balance Principle. John Wiley & Sons, New York. Baumgärtner, S., 2000. Ambivalent Joint Production and the Natural Environment. An Economic and Thermodynamic Analysis. Physica-Verlag, Heidelberg, New York. Baumgärtner, S., H. Dyckhoff, M. Faber, J.L.R. Proops, and J. Schiller, 2001. The concept of joint production and ecological economics. Ecological Economics 36:365–72. Baumgärtner, S., M. Faber, and J. Schiller, 2006. Joint Production and Responsibility in Ecological Economics – On the Foundations of Environmental Policy. Edward Elgar, Cheltenham, UK, and Northampton, MA. Boulding, K., 1966. The economics of the coming spaceship Earth. In: H. Jarrett (ed), Environmental Quality in a Growing Economy. Johns Hopkins University Press, Baltimore, pp. 3–14. Christensen, P.P., 1989. Historical roots for ecological economics – Biophysical versus allocative approaches. Ecological Economics, 1:17–36. Cornes, R., and T. Sandler, 1984. Easy riders, joint production, and public goods. The Economic Journal, 94:580–98. Daly, H.E., 1977. Steady State Economics: The Economics of Biophysical Equilibrium and Moral Growth. W. H. Freeman, San Francisco. Debreu, G., 1951. The coefficient of resource utilization. Econometrica, 19:273–92. Debreu, G., 1959. Theory of Value. An Axiomatic Analysis of Economic Equilibrium. John Wiley & Sons, New York. Dyckhoff, H., 1996. Kuppelproduktion und Umwelt. Zur Bedeutung eines in der Ökonomik vernachlässigten Phänomens für die

Kreislaufwirtschaft. Zeitschrift für angewandte Umweltforschung, 9:173–87. Faber, M., R. Manstetten, and J.L.R. Proops, 1992. Humankind and the environment: an anatomy of surprise and ignorance. Environmental Values, 1:217–42. Faber, M., H. Niemes, and G. Stephan, 1983/1995. Entropie, Umweltschutz und Rohstoffverbrauch: Eine naturwissenschaftlich ökonomische Untersuchung [Entropy, Environment and Resources: An Essay in Physico-Economics]. Spinger-Verlag, Heidelberg. Faber, M., J.L.R. Proops, and S. Baumgärtner, 1998. All production is joint production – a thermodynamic analysis. In: S. Faucheux, J. Gowdy, and I. Nicolaï (eds), Sustainability and Firms. Technological Change and the Changing Regulatory Environment. Edward Elgar, Cheltenham, UK, and Northampton, MA, pp. 131–58. Georgescu-Roegen, N., 1971. The Entropy Law and the Economic Process. Harvard University Press, Cambridge, MA. Koopmans, T.C., 1951. Analysis of production as an efficient combination of activities. In: T.C. Koopmans (ed), Activity Analysis of Production and Allocation. John Wiley & Sons, New York, pp. 33–97. Kurz, H.D., 1986. Classical and early neoclassical economists on joint production. Metroeconomica, 38:1–37. Lindahl, E., 1919. Die Gerechtigkeit der Besteuerung. Gleerup, Lund. Mas-Colell, A., M.D. Whinston, and J.R. Green, 1995. Microeconomic Theory. Oxford University Press, New York. McKenzie, L.W., 1959. On the existence of general equilibrium for a competitive market, Econometrica, 27:54–71. Nadiri, M.I., 1987. Joint production. In: M. Vernengo, E. Perez Caldentey, and B.J. Rosser, Jr. (eds), The New Palgrave Dictionary of Economics. Palgrave Macmillan, London, pp.  1–5. https://​doi​.org/​10​.1057/​978​-1​-349​ -95121​-5​_731​-1 Norton, B.G., 1992. Ecological health and sustainable resource management. In: R. Costanza (ed), Ecological Economics. The Science and Management of Sustainability. Columbia University Press, New York, pp. 102–17. O’Connor, M., 1993. Entropic irreversibility and uncontrolled technological change in economy and environment. Journal of Evolutionary Economics, 3:285–315. Oenning, A., 1997. Theorie betrieblicher Kuppelproduktion. Physica-Verlag, Heidelberg. Perrings, C., 1994. Conservation of mass and the time-behaviour of ecological-economic systems. In: P. Burley and J. Foster (eds), Economics and Thermodynamics: New

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nuclear fuel cycle: an historical and discourse analysis. Ecological Economics, 39:13–19. Spengler, T., 1999. Industrielles Stoffstrommanagement. Erich Schmidt Verlag, Berlin.

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56. Kapp, Karl William Born in 1910 in Königsberg, Germany, Karl William Kapp graduated in economics and law at the University of Berlin. In 1933, to escape the Nazi persecutions, he moved to Geneva where he obtained a PhD discussing a thesis on ‘economic planning and international commerce’. After lecturing at various universities in the US from 1937 to 1957, he spent several periods as a visiting professor in India and the Philippines, where he was able to observe the scarce success of traditional development policies elaborated without considering their cultural and institutional contexts. In 1965 he returned to Switzerland as a professor at the University of Basel. He acted as a consultant for the first United Nations Conference on the Human Environment held in Stockholm in 1972. He died from a heart attack while participating in a conference on ecological development at the University of Dubrovnik in Croatia on 10 April 1976. Interested readers will find more details in a biography written by Kapp’s last assistant, Rolf Steppacher (1994), and on the website dedicated to him (see www​.kwilliam​ -kapp​.de/​who​.htm). Well ahead of his time, in 1950 he published a book reporting many sources of ‘social costs’ generated by the competitive capitalist economy (Kapp, 1950).1 His thoughts came to a mature expression in several articles published mainly in the journal Kyklos from the 1960s to his premature death. He belonged to the old-institutionalist school of thought, influenced in particular by T. Veblen, J.M. Clark, G. Myrdal, A. Lowe, F. Perroux, and K. Polanyi. At the same time, he was also aware of the major developments occurring in the late 1950s and 1960s in systems thinking and cybernetics (e.g. Ackoff, 1960, and von Bertanlaffy, 1968). Kapp masterfully combined ingredients from these two fields into a unitary framework that anticipated most of the elements that would later become the core of ecological economics, as also shown by Røpke (2004). Two other entries in this Encyclopedia largely refer to specific aspects of Kapp’s thought, namely, ‘The Economy as an Open System’, entry 26, and ‘Cost Shifting and the Competitive Society’, entry 15. The present entry focuses on his ideas as a whole. Its purpose is to place them in an interpretative

framework able to show the unitarity of the different aspects of Kapp’s thought, whose links might otherwise be difficult to see. This allows us to highlight the fact that his key arguments logically follow from viewing the economy as an open biophysical system: Kapp (1976) himself stressed how difficult it is to understand the many epistemological implications of the rather obvious statement that the economy is an open system. Finally, a unitary framework makes it easier to draw comparisons with the ideas that found ecological economics. As in Georgescu Roegen, whose work was cited by Kapp (e.g. Kapp 1977), societal metabolism is central. The economy and society exchange matter with the natural environment, similarly to what living beings do to keep themselves far removed from the thermodynamic equilibrium, that is, alive (Kapp, 1961, 93). Kapp stressed the importance of societal metabolism for environmental disruption, stating that the key problem of the open-system character of the economy … [is] that production derives material inputs from the physical and decisive impulses from the social system which, in turn, may be disrupted and disorganised by the emission of residual wastes up to a point where social reproduction itself may be threatened. (Kapp, 1976, 98; see also Kapp, 1977, 531–2)

Addressing the biophysical dimension involves acknowledging the importance of system dynamics and the embeddedness of the economy within the social sphere, both embedded in the natural environment. This is why matter (and energy) is at the base of the tree in Figure 56.1, summarising Kapp’s main ideas, complemented on the left and right, respectively, by a sketch of a system and of embeddedness. Two epistemological arguments logically follow: on the one hand, we must admit that phenomena unravel on many spatio-temporal scales (e.g. Kapp, 1976, 99; Kapp, 1977, 529 ff); on the other, only interdisciplinary research can successfully approach the problem of environmental disruption (e.g. Kapp, 1977, 528). Both cornerstones – respectively, the need for several non-equivalent descriptions of the same phenomenon (Giampietro, 2003), and inter-disciplinarity (Costanza 1989) or trans-disciplinarity (e.g. Baumgärtner et al. 2008) – then became keystones in ecological

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economics. Then four interconnected pillars follow (system dynamics, change, the already mentioned pervasiveness of social costs, and incommensurability) around which the main elements of Kapp’s thinking can be framed.

and Spash and Fuselier, entry 15 in this Encyclopedia). Moving to the right in Figure 56.1, the next branch shows the importance of change in Kapp’s thought, which also follows from

Figure 56.1 Kapp in a snapshot: the economy as an open system and its implications

As mentioned above, system theory became increasingly topical between the 1950s and 1960s, greatly affecting the development of ecology as well. Kapp perceived that the dynamic equilibria and changes in a system are the outcome of negative and positive feedback loops linking its elements and regulating its functioning, and that non-linearities are often at play, determining thresholds and synergies (e.g. Kapp, 1976, 97 ff, or Kapp, 1977, 529 ff). The left branch of the tree in Figure 56.1 shows a series of concepts from system theory that Kapp used in his analysis. They include ‘circular cumulative causation’, an expression used by Myrdal to indicate ‘positive feedback loops’ that was further elaborated by Kapp (see Luzzati, 2009, 2010,

his adherence to system theory and the old-institutional approach. Unlike neoclassical economics, Kapp (1976, 102) insisted on the importance of considering change not only in technology and institutions, but also in individual preferences. Following Mill and Veblen, Kapp took individuals not as passive automata choosing available means that best fit given aims, but rather as active, learning and social beings. For him, ‘man, with his specifically human intelligence, is capable of using reason and science for the exploration of goals and as a basis for judgements as to the kind and direction of action to be followed’ (Kapp, 1965, pp. 76–7). The next key concepts shown in Figure 56.1 (second branch from the right) are Tommaso Luzzati

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incommensurability and the need for multidimensional evaluations in social choices. Cornerstones in ecological economics (see Martinez-Alier et al., 1998), they follow straight from accepting complexity, as stated clearly by Kapp: ‘the heterogeneous character of the disrupting flows of damages and the complex interdependencies to which we have referred above preclude any measurement and evaluation in terms of a common denominator’ (Kapp, 1970, p. 846). Arguing in favour of an integrated multidimensional assessment (Kapp, 1976, 97), he wrote that ‘all monetary evaluations … [are] problematical if not indeed unacceptable and cognitively irrelevant’ (Kapp, 1976, 101). This applies particularly to collective decisions for which monetary evaluation is unable to express the ‘relative social importance in the sense of value to society (and individuals) both in the short and in the long run’ (Kapp, 1976, 101) of environmental damage and of public goods and services. Real-world complexity cannot be reduced to a single dimension (see Kapp, 1977, 534), expressing serious doubts regarding the ‘current proposal of “deducting” social costs from gross or net national product measurements’ (Kapp, 1976, 104). As for Georgescu Roegen, for Kapp, synthetic measurements ‘upon closer analysis, can be shown to reflect either the subjective preferences and valuations of the experts and/or powerful vested interests’ (Kapp, 1976, 100). To refute incommensurability and monetary evaluation is not an impediment for public decisions. ‘The elaboration and acceptance of environmental goals call for a collective or social choice with direct participation and expression of preferences by all members of society, even those outside the market and without reference to effective demand’ (Kapp, 1963, p. 317). In other words, the many conflicting interests and perspectives have to be reconciled through democratic processes2 (e.g. Kapp, 1976, 100; or Kapp, 1977, 536–7). To this end, a broad set of social and environmental indicators must be available, which was lacking at that time, as often decried by Kapp (e.g. 1974b). Finally, also the goals of economic policy should be radically different in Kapp’s view from those of mainstream economics. For Kapp, top priority has to be given to ‘the social and moral imperative of minimizing human suffering’ (Kapp, 1977, 538) and to Tommaso Luzzati

ensure human well-being and even survival, which are threatened, through environmental disruption, by the economic process (Kapp, 1976, 91). In other words, economic and development policy should prioritise satisfaction of basic needs and compliance with environmental limits (Kapp, 1976, 101). A decade later, such a perspective was included in the United Nations Brundtland report which, just after defining it, specified that sustainable development ‘contains within it two key concepts: the concept of “needs”, in particular the essential needs of the world’s poor, to which overriding priority should be given; and the idea of limitations imposed by the state of technology and social organization on the environment’s ability to meet present and future needs’ (Brundtland, 1987, 41). Hence, in Kapp’s humanistic view, compromise between the conflicting interests cannot exceed the limits imposed by ethics and by what science suggests about the maintenance of dynamic states of the ecosystems. To sum up, ‘to satisfy these human needs … [the environmental] requirements will have been defined as objectively as our present knowledge permits and evaluated by means of a deliberate collective, i.e. political decision in comparison to other public goals to be pursued’ (Kapp, 1963, 317). Finally, the first branch from the right recalls the key concept of social costs that Kapp started elaborating in his book in 1950. Despite some similarity with the notion of externality, it is much wider and analytically more useful (e.g. Berger, 2015, 2017). Acknowledging the material dimension of the economic process, external effects must be seen as ubiquitous rather than special cases, as in neoclassical economics. Moreover, the systemic pressure arising from economic competition makes it impossible to consider externalities as unintended side-effects; on the contrary, they are generated for the purpose of ‘cost shifting’ (see Spash and Fuselier, entry 15 in this Encyclopedia). As a consequence, one of Kapp’s main arguments was that, if unregulated, ‘the organising principles of economic systems guided by exchange values are incompatible with the requirements of ecological systems and the satisfaction of basic human needs’ (Kapp, 1976, 95). Tommaso Luzzati

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Notes

Kapp, K.W., 1970. Environmental disruption and social costs: a challenge to economics. Kyklos, 23(4), 833–48. Kapp, K.W., 1974a. The implementation of environmental policies, in K.W. Kapp (ed), Environmental Policies and Development Planning in Contemporary China, and Other Essays, pp. 101–26. Mouton. Kapp, K.W., 1974b. Environmental indicators as indicators of social use value, K.W. Kapp (ed), Environmental Policies and Development Planning in Contemporary China, and Other Essays, pp. 127–38. Mouton. Kapp, K.W., 1974c. Environmental Policies and Development Planning in Contemporary China, and Other Essays. Mouton. References Kapp, K.W., 1976. The open system character of Ackoff, R.L., 1960. Systems, Organizations and the economy and its implications, in K. Dopfer Interdisciplinary Research. General Systems (ed), Economics in the Future: Towards a New Yearbook, 5, 1–8. Paradigm, pp. 90–105. Macmillan. Baumgärtner, S., Becker, C., Frank, K., Müller, Kapp, K.W., 1977. Environment and technolB., & Quaas, M., 2008. Relating the philosogy: New frontiers for the social and natural ophy and practice of ecological economics: sciences. Journal of Economic Issues, 11(3), The role of concepts, models, and case studies 527–40. in inter-and transdisciplinary sustainability Luzzati, T., 2009/2014. Human needs, sustainable research. Ecological Economics, 67(3), 384–93. development and public policy: learning from Berger, S., 2015. K. William Kapp’s social theory KW Kapp (1910–1976). In N. Salvadori & of social costs. History of Political Economy, A. Opocher (eds), Long-Run Growth, Social 47(S1), 227–52. Institutions and Living Standards, pp. 305–22. Berger, S., 2017. The Social Costs of Neoliberalism Edward Elgar Publishing. – Essays on the Economics of K. William Kapp. Luzzati, T., 2010. Economic development, Spokesman. environment and society: rediscovering Karl Brundtland, G.H., 1987. Our Common Future: William Kapp. In J.R. McNeill, J.A. Pádua, & Report of the World Commission on M. Rangarajan (eds), Environmental History: Environment and Development. https://​ As If Nature Existed, pp.  48–64. Oxford su​stainabled​evelopment​.un​.org/​content/​ University Press. documents/​5987our​-common​-future​.pdf Martinez-Alier, J., Munda, G., & O’Neill, J., Costanza, R., 1989. What is ecological econom1998. Weak comparability of values as a founics? Ecological Economics, 1(1), 1–7. dation for ecological economics. Ecological Giampietro, M., 2003. Multi-Scale Integrated Economics, 26(3), 277–86. Analysis of Agro-Ecosystems. CRC Press. Røpke, I., 2004. The early history of modern Kapp, K.W., 1950. The Social Cost of Private ecological economics. Ecological Economics, Enterprise. Harvard University Press. 50(3–4), 293–314. Kapp, K.W., 1961. Toward a Science of Man in Steppacher, R. 1994. ‘Kapp, K. William’, in Society. A Positive Approach to the Integration The Elgar Companion to Institutional and of Social Knowledge. Martinus Nijhoff. Evolutionary Economics, Aldershot, UK and Kapp, K.W. 1963/1978. The Social Costs of Brookfield, US: Edward Elgar, pp. 435–41. Business Enterprise. Spokesman. von Bertalanffy, L., 1968. General System Theory. Kapp, K.W., 1965. Economic development in George Braziller. a new perspective: existential minima and substantive rationality. Kyklos, 17(1), 49–79. 1.

The second edition (1963) broadened and deepened the arguments. The title was changed to Social costs of business enterprise to emphasise that public businesses generate social costs as well. In 1971, the 1950s book was reprinted (by Schocken) with the addition of a very interesting introduction. 2. On democratic processes, Kapp (1974a) wrote: ‘The author has no illusions about the fact that such a transformation will come about by itself and without struggle. It calls for a genuine democratisation of the state (that is to say, of the centre of political power) and of the economy at all levels, i.e. at the micro level of the firm, the regional and the central level of policy-making’ (p. 138).

Tommaso Luzzati

57. Land grabbing Introduction

Drivers and dimensions of land grabbing

Land grabbing refers to the recent and unprecedent changes in control over and access to land that have occurred globally since the onset of the 21st century (Borras et al., 2011). The term was originally coined by the civil society organization GRAIN (2008) to denounce colonial-like, transnational, large-scale, and long-term farmland acquisitions announced by international investors and states in pursuit of food security and new investment opportunities, provoking vast concerns over dispossession and displacement of customary land users. Current debates on land grabbing address a wide range of projects and processes beyond farmland acquisitions that involve a variety of actors. Other related terms frequently used in the literature to describe the global land grab phenomenon are large-scale land acquisitions (LSLAs), land deals, or the global land rush. Land acquisitions, land speculation, agricultural expansion, and associated changes in effective control over and access to land are nothing new in human history. However, the land grab phenomenon has been characterized by unprecedented growth in demand for land globally. Land grabs spiked and made it to the news headlines worldwide following the global 2007–08 food price crisis and a resulting farmland investment boom (Fairbairn, 2014). While before 2008, annual agricultural land expansion was below 4 million ha, the limited information available on land deals announced during 2008–09 indicated a demand of 45 million ha globally (World Bank, 2010). A report by the International Land Coalition stated that between 2000 and 2011 more than 200 million ha of land had been subject to negotiations for land acquisition, with Africa being the primary target (Anseeuw et al., 2012). More recent estimates indicate that since the year 2000 at least 160 million hectares have been under negotiation for lease and purchase (Land Matrix, 2021). Such estimates depend strongly on the definitions of land grabbing employed in these studies and are limited by the information available in large databases. According to Borras (2022), the actual scope and extent of land grabbing globally may be much wider.

Agricultural investment and associated financial and land speculation have been important causes of land grabs. However, the actual range of underlying drivers is diverse (Arezki et al., 2015; Scheidel and Sorman, 2012; Zoomers, 2010). Motives for land grabbing can include land acquisitions from customary groups and local communities for infrastructure development, such as special economic zones, residencies, tourism complexes, transport infrastructure, mining concession, hydropower dams, and other energy infrastructures, as well as land acquisitions and exclusions from land resources to set up conservation areas and programs, forestry concessions, or reforestation and afforestation projects, among other motives. Land grabs justified or implemented for environmental ends have been specifically discussed under the term green grabbing (Fairhead et al., 2012). The enormous interest from academics, activists, civil society, and governmental actors to better understand and address the land grab phenomenon, as well as the ample application of the term to study a large variety of processes involving changes in land control and access, has produced a substantial body of literature that some have described as the land grab literature rush (Oya, 2013). Scholars linked to the Land Deal Politics Initiative, an international network of academics concerned with global land grabbing, published several edited collections that offer diverse perspectives on the global land grab (Borras et al., 2011), and discuss its various dimensions, such as the role of biofuels in reshaping land control (Borras et al., 2010), the role of the state for land governance (Wolford et al., 2013), the relation of global land grabs to globalization and shifts in global capital accumulation (Margulis et al., 2013), the bottom-up political reactions from affected groups (Hall et al., 2015), or the methods and epistemologies used in tracing and analysing the global land grab phenomenon (Scoones et al., 2013). In addition to numerous academic case studies and civil society reports, several online platforms register land deals and related land acquisition conflicts systematically, most notably the Land Matrix (Nolte et al., 2016) and the Environmental Justice Atlas (Scheidel et al., 2020).

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Socio-ecological implications of land grabbing

The socio-ecological impacts and implications of land grabs vary in relation to the case-specific characteristics, that is, the land use history and the previously existing social and environmental vulnerabilities at play at a given location, the purpose of the acquisition and related resource demands, involved actors, as well as procedural and distributional aspects, among other factors. However, systematic reviews and multi-case analyses have identified severe adverse impact trends provoked by land grabbing that threaten ecological sustainability, livelihoods, and social justice. 403 A multi-country assessment of 82  land deals found direct and indirect land use change provoked by land grabs to be a significant driver of tropical forest loss (Davis et al., 2020). Related land use changes are expected to have high fossil-energy footprints because of the use of fertilizers and the irrigation and mechanization required for the development of commercial agriculture, while adversely affecting local populations’ access to energy resources (Rosa et al., 2021). Consequently, land grabs are estimated to increase carbon emissions globally through deforestation (Liao et al., 2021) and the further industrialization of agriculture (Rosa et al., 2021). Land grabs also entail large water appropriations (Breu et al., 2016; Chiarelli et al., 2016) which in turn affect local populations’ access to water and the right to food (Dell’Angelo et al., 2018). Furthermore, land grabs are a potential driver of landslides and soil erosion (Chiarelli et al., 2021), affect natural habitats and their fragmentation, and jeopardize biodiversity (Debonne et al., 2019). In sum, the global land grab phenomenon is expected to provoke lasting global environmental change (Lazarus 2014). The socio-economic impacts of land grabs are far-reaching. Early policy papers from mainstream development institutions argued that the rise of investment into farmland could potentially bring benefits for host countries and local populations, particularly in terms of raising food production capacities, creating new employment opportunities and incomes, and developing rural infrastructure (Cotula et al., 2009; World Bank, 2010). However, multi-case analyses showed that by re-orienting crop production to nutrient-poor

crops predominantly destined for export, and by limiting local populations’ access to land, land deals produce significant food security risks (Müller et al., 2021). Other studies pointed to the opportunity costs of employment creation and poverty reduction that arise form allocating farmland to agribusiness instead of small-farmer enterprises (De Schutter, 2011; Scheidel et al., 2013). Numerous case and systematic review studies documented adverse livelihoods impacts on the ground, arising through enclosure of livelihood assets and elite capture (Oberlack et al., 2016), crowding out of small-farmers (Nolte and Ostermeier, 2017), loss of environmental incomes (Jiao et al., 2015), and reducing local populations’ access to commons managed under customary institutions (Dell’Angelo et al., 2017a). Vulnerable segments of society are particularly affected by land grabbing and growing resource insecurity. Indigenous Peoples, who are increasingly confronted with encroachment of commons, loss of traditional custodianship of their lands, dispossession, and displacement, face disproportionately high levels of violence in land acquisition conflicts (Dell’Angelo et al., 2021; Scheidel et al., 2023). Furthermore, the social burdens of land deals tend to be gendered and unevenly distributed among household members (Atuoye et al., 2021; Nyantakyi-Frimpong and Bezner Kerr, 2017). In general, land deals raise important social justice questions regarding access to land in the context of global agrarian and environmental change and the looming climate crisis (Borras and Franco, 2018; Borras et al., 2020; Franco et al., 2017; Hunsberger et al., 2017).

Land grabbing and climate change

Land grabbing intersects with climate change mitigation and adaptation politics in various ways. One the one hand, land-based climate change policies have substantially altered the politics of land access, acting as an important driver of land grabbing (Franco and Borras, 2019; Hunsberger et al., 2017). Examples are land acquisitions linked directly and indirectly to the expansion of biofuel crops plantations (e.g. Aha and Ayitey, 2017), Afforestation and Reforestation (A/R) projects (e.g. Scheidel and Work, 2018), REDD+ (Ingalls et al., 2018), conservation areas (Schleicher et al., 2019), renewable energy Arnim Scheidel

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installations (e.g. Sovacool, 2021), or natural disaster management (e.g. Uson, 2017). On the other hand, land grabbing jeopardizes both climate change mitigation and adaptation processes. Land grabs have become an important source of greenhouse gas emissions globally (Liao et al., 2021; Rosa et al., 2021). Global land grabbing directly affects the vulnerability of local populations and their capacity to adapt to a changing environment, bearing important implications for climate change adaptation. Populations displaced by land deals may face higher exposure to climate change impacts (Dell’Angelo et al., 2017b). Local communities’ declining access to livelihood resources and deteriorating food security and sovereignty exacerbates previous and creates new, gendered vulnerabilities (Atuoye et al., 2021; Yengoh et al., 2015). Land grabs affecting common-pool resources governed by Indigenous and customary institutions jeopardize the resilience and adaptive capacity of local socio-ecological systems (Dell’Angelo et al., 2017a; Haller et al., 2020). Growing land tenure insecurity may force farmers to engage in unsustainable farming practices (Aha and Ayitey, 2017), impede community-based sustainable forestry (Gabay and Alam, 2017), and hinder agroecological innovations to manage climate risks (Nyantakyi-Frimpong, 2020).

Pathways forward

Avoiding, addressing, and reversing increased vulnerabilities provoked by land grabbing requires that land governance includes the principles of land recognition, land redistribution, and land restitution (Borras and Franco, 2018). Land recognition is necessary to avoid land acquisitions in the first place, particularly for lands used and governed by Indigenous and ethnic minorities. Land redistribution increases the asset base of local populations and may enable climate resilient agro-food systems. Land restitution is needed for those who have been displaced, facing particularly high vulnerability to socio-ecological crises, including exposure to climate change impacts. Future research that aims to better understand and support processes of land recognition, redistribution, and restitution may investigate the potentials and pitfalls of successful and failed cases worldwide, look into Arnim Scheidel

the conditions and resistance strategies under which local communities have been able to effectively reclaim their lands and engage in the co-production of knowledge, methods, and evidences relevant to support agrarian and environmental justice within a framework of scholar activism (Borras, 2016). Such knowledge co-production requires a collaborative action-research approach capable of accounting for local understandings of justice, as well as past and present histories and practices of land use threatened and transformed by land grabbing (Hunsberger et al., 2017).

Acknowledgements

Parts of this entry were initially prepared for the 6th IPCC report, working group III contribution on Impacts, Adaptation and Vulnerability, and later adapted for this encyclopedia entry. The author acknowledges financial support from the Spanish Ministry of Science, Ramón y Cajal fellowship program (RYC2020-029088-I). Arnim Scheidel

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Canadian Journal of Development Studies 38, 414–30. https://​doi​.org/​10​.1080/​02255189​ .2017​.1308316 Wolford, W., Borras, S.M., Hall, R., Scoones, I., White, B., 2013. Governing global land deals: the role of the state in the rush for land. Development and Change 44, 189–210. https://​ doi​.org/​10​.1002/​9781118688229 World Bank, 2010. Rising Global Interest in Farmland: Can It Yield Sustainable and Equitable Benefits? Yengoh, G.T., Armah, F.A., Steen, K., 2015. Women’s bigger burden: disparities in outcomes of large scale land acquisition in Sierra Leone. Gender Issues 32, 221–44. https://​doi​ .org/​10​.1007/​s12147​-015​-9140​-7 Zoomers, A., 2010. Globalisation and the foreignisation of space: seven processes driving the current global land grab. Journal of Peasant Studies 37, 429–47.

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58. Land-time budget analysis Land-Time Budget Analysis (LTBA) is a method used to analyse the parallel investment of two of the most essential resources applied by smallholder farmers: labour (available personal time) and land resources (available area, potentially used in agriculture and other land uses). The method was developed by Mario Giampietro and others around the turn of the millennium, initially to analyse farmer household types in Asia (Pastore et al. 1999; Gomiero & Giampietro 2001; Giampietro 2004). LTBA is an analytical method that (i) represents a parallel reading of two resource use profiles by avoiding reductionism; (ii) develops an understanding of resource investment decisions faced by smallholders, as well as the social and biophysical constraints they operate under; (iii) yields a sequence of indicators that allow users to construct smallholder household types, which can be easily communicated to policy actors; and (iv) can be translated into monetary terms to gain accurate and robust productivity indicators of land and labour. Major applications of LTBA include understanding local resource use decisions, the sustainability of livelihood systems, and developing scenarios for different household resource use types. Time and land use analysis goes back to agricultural studies during the latter half of the 1900s. These studies conceptualised time as a major resource for farming folk before the advent of heavy mechanisation and industrial agriculture. Understanding that productivity indicators, expressed in money units per resource unit invested, have limited bearing on smallholder decisions, social scientists began to look at decisions surrounding the use of biophysical resources with the emergence of the social metabolism concept (Georgescu-Roegen 1971). While material and energy use were conceivable within this framework, time resources were initially disregarded as a resource per se; rather, the energy expended as a function of labour was initially used as an input to production (Rappaport 1969). Time, as a physical constant divided into human-made units, however, was an oddity within a biophysical accounting framework. Nevertheless, social

scientists, such as Carlstein (1982), Groh (1992), and Geshuny (2000), considered time a major constraint in the production and consumption cycles of societies, regardless of whether they were non-industrial or modern capitalist. Land use studies, on the other hand, are well established in ecological and agricultural studies, and land is well recognised as a primary resource to agriculturalists, regardless of the epoch (Boserup 2002). Linking land use with agricultural technology, food security, productivity, and sustainability has been well established among students of cultural evolution, agronomy, adoption studies, and international development research. LTBA, however, brought together two types of analysis and maintains that both land use and the intensity and amount of effort invested into it are relevant to the development outcomes of any community (cf. Carlstein 1982). In the absence of capital and limited access to technology, time and land are essential resources in subsistence economies.1 Today, semi-subsistent farming communities can still be found across the world and on every continent. In fact, smallholders – small-scale farmers with some degree of self-sufficiency – still make up the bulk of the globe’s agricultural population (Food and Agriculture Organization [FAO] 2012). In taking critical decisions concerning their livelihood strategies, they face three general choices among alternative economic objectives: (1) minimisation of risk and maximisation of food and livelihood security; (2) maximisation of productivity and profits; and (3) maximisation of leisure, or non-productive (reproductive) time. Each farming household can be profiled against these three sets of preferences, and each of these considerations implicitly or explicitly influences local-level farmer decisions. These priorities are mediated by cultural norms, social institutions (such as peer pressure, local laws, leadership), and individual preferences, as much as they are constrained by external influences, such as ecological conditions, markets, and taxation. Generally, however, household profiles will display a mix of these strategies, although non-industrial societies tend to err on the side of risk minimisation and maximisation of leisure (Mayrhofer-Grünbühel 2004).

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Theoretical foundations

Sahlins (1974) described the “original affluent society” among hunter-gatherers who needed only 3–4 hours per day to fulfil basic needs (food, housing, and protection against the elements and wild animals). Upon fulfilling livelihood basics, the rest of the day is spent in “leisure”. However, leisure in non-industrial societies is drastically different than the work–play dichotomy in modern industrial society. Non-productive time is spent in (i) social communication; (ii) education and care; (iii) ritual; and (iv) information exchange. In egalitarian societies, and in the absence of state-level institutions, reproduction of social institutions is critical. This is achieved through social communication – communicating norms, rules, beliefs, and preferences – as well as through ritual: establishing the group identity and affirming corresponding hierarchies of power. Education and care – in subsistence societies, usually not outsourced to professionals – are part of the social reproduction process necessary to sustain the community. Cultural evolution and innovation come from a process of information exchange in which information is accessed to form decisions on various aspects of life, including livelihood strategies (Groh 1992). Risk minimisation ensures the survival of the community and prevents “collapse” from occurring. Two main strategies are remarkable here: (1) underproduction (cf. also entry on Economic Anthropology in this volume); and (2) diversification. Underproduction refers to the tendency of subsistence societies to remain well beyond the carrying capacity of a given ecological unit. Not all resources are extracted from a given area, but some are left for ecological reproduction, as a “reserve” in times of resource scarcity or hardship (Boserup 2002). By remaining within ecological limits, sustainability is ensured, and resilience increased. Underproduction is a built-in safety feature and the way “redundant systems” cope with uncertainty (Giampietro 2004). Diversification, next to underproduction, is the second remarkable feature of risk-minimising farming communities (Ellen 1996). Diversification occurs at various levels: spatially, by spreading out production areas across available land, rather than concentrating production in one spot; func-

tionally, by diversifying livelihood activities (food crops, cash crops, wage labour, handicrafts); ecologically, by growing diverse crops and diverse varieties within each crop type (Boserup 2002). Modern societies, on the other hand, are also highly diversified, although at the societal level, not at the household level. They are not redundant systems, but highly specialised, vulnerable to external change. Systematic underproduction is of no value as profit maximisation becomes the top preference, including at the expense of leisure; or “leisure” becomes more work-like: consumption of services instead of reproduction of social institutions. According to Gershuny (2000), as societies become more differentiated, so does the use of time resources: instead of shared experiences for the entire group, work and leisure shift to certain social groups and time is consumed differently according to each group within the same community. In general, however, there is a marked shift towards paid work due to three factors: (a) a growing economy means more people spending more time for paid work; (b) more women joining the formal workforce over the last 70 years; (c) leisure becomes consumption and unproductive work (“annual leave”) decreases.2 Modern societies are, thus, faced with a dilemma: while the production of goods becomes more productive in terms of output per time invested, more (time) needs to be consumed to absorb the increased production capacity. At the same time, goods become more sophisticated and production processes are continuously added to produce the same goods (“adding value”), so that more people are required in the workforce. Productivity increases are offset by higher complexity of production processes. This speaks to the “Jevons paradox” described by Giampietro (2004: 7) in that “an increase in efficiency in using a resource leads, in the medium to long term, to an increased use of that resource (rather than to a reduction)” – in this case, time use. On the other hand, the least sophisticated farming systems (= low-input, low-mechanised agriculture) require the least labour in productive activities related to the provisioning of food and basic needs. Increased labour demand corresponds with the gradient observed from subsistence farming to higher degrees of market integration (cf. Sahlins 1974). Clemens M. Grünbühel

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Based on ecological principles, Carlstein (1982), following Odum (1971), defines energy and “living-space” as any society’s main resources, as they must use these to render all other resources useful when intervening in nature. However, how these resources are applied are a matter of cultural and institutional organisation, as well as the technology available. The process of resource conversion is termed “work”, while all other processes – still using energy and space – are non-work. Even though energy and space are indispensable, they are limited and, thus, constraining. While a society’s time budget is limited by the number of people in the group, land is limited by the area a particular group has access to. Any area occupied by people is, thus, limited by space-time, expressed, for example, in hectare-days, which, in turn, defines the carrying capacity of that area (Carlstein 1982). In non-industrial societies, space-time levels depend on available technology, as well as the duration and distance people are able or willing to travel. For example, in shifting cultivation, the system works well and is sustainable if the spread of plots is not restrained by geographical, legal, or political factors – that is, the community has access to pristine or mature-growth forest and can allow sufficient time for old plots to re-grow. Issues of ecological and social nature arise, however, when population expansion, the state, and competing communities restrict access, which then leads to a gradual degradation of the system, originally well balanced-out across space-time (Carlstein 1982). Time investments can be made into a variety of activities and in a variety of locations. They are dispersed according to individual or group choices. By making these decisions (the when and the where), a group essentially adapts to their environment using the available technology (Carlstein 1982). While in today’s modern financial economy time and space may play a minimal role or no role at all, in agriculture these are the critical dimensions (Gershuny 2000). Capital is essentially used to access technology or inputs. While capital may allow access to time-saving machinery, the time gains are usually offset by higher energy expenditures, thereby reducing overall energy efficiency; for example, harvesting by hand takes more time but is overall more energy efficient than using a combine harvester, which Clemens M. Grünbühel

requires diesel for combustion (Smil 2017).3 Regardless, modern technology is usually limited in semi-subsistent societies, thus, access to it is unavailable despite attempts by development agencies. The majority of the world’s poor remain in rural areas, in some way or another connected to agriculture (FAO 2012). Low-input, low-mechanisation systems, thus, bind labour and take place in a limited land area (Grünbühel 2016). LTBA analyses the allocation of these two resources, looks at the choices the community makes regarding them, and relates the decisions on land and time use so they become congruent with each other. LTBA is scalable – that is, it can be (and has been) applied to scales ranging from households to national economies (Gomiero & Giampietro 2001; Grünbühel & Schandl 2005; Serrano-Tovar & Giampietro 2014). Additionally, LTBA yields several system performance indicators, such as productivity, resilience, and carrying capacity, which support an understanding of any given agrarian social group in ways that traditional productivity indicators or performance averages have not been able to deliver before (Giampietro 2004).

Methods

Data are gathered using common field research methods, such as structured observation, time-diary methods, or quantitative/ semi-structured interviews. Alternatively, time use statistics are available in several countries, sometimes at subnational levels. In essence, the method implies an overall time budget, calculated as the number of people multiplied by their available total time (number of days/hours/minutes in a given time period). From this time budget, a “physiological overhead” is subtracted (i.e. maintenance of our bodies, which every individual needs to conduct to physiologically survive). This includes mostly sleep and basic hygiene (culturally variable). In many societies, the physiological overhead can reach up to 50 per cent of the overall time budget. The remainder is disposable time – essentially the time available to humans, which they can decide over how to invest.4 The first decision is on “productive” versus non-productive activities. In Giampietro’s (2004) terms, this is “social overhead” (i.e. the time required to reproduce social institutions within the com-

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munity, including education, care, rituals, mating, and recreation), and “available work time” (human activity in work, HAwork in Figure 58.1). The latter is the disposable time in productive activities invested in either subsistence (informal; food- and needs-related; non-monetised; Wsub in Figure 58.1) or monetised work (W$ in Figure 58.1). Monetised work is then further subdivided into direct sale of labour (W-Offarm) or cash cropping (W-land). This cascade of divisions provides an overview of various types of human activity and their purpose; on the other hand, it also displays how much labour is actually used for formal income rather than in reproductive and subsistence activities.

Source:

resources are extracted, however in- or extensively. This land can be divided into land not in agricultural production (LNAP) and land in production (LIP), on which agriculture occurs. The former may consist of forest, fallow, or degraded land used for hunting and gathering of non-timber forest products. LIP, in turn, is further subdivided into land for basic needs (LIPsub) and land for cash crops (LIP$). The latter can again be separated into land needed to pay for agricultural inputs and taxes, while the former is land that provides income to the farming household. As a validity check, the monetary value derived from L-NDC (Land for net disposable cash) must equal the income derived from W-land (work

Giampietro (2004, Figure 11.8, p. 396).

Figure 58.1 Schematic of Giampietro’s (2004) Land-Time Budget Analysis, representing farmer decisions, into which activities and areas are invested

In analogy to investments in time, land use can be represented similarly. The land budget (i.e. the total available land [TAL]; cf. Figure 58.1) is initially divided into the “ecological overhead”, or land not in production, either because it is unusable or protected for whatever reason (e.g. religion, conservation), and “colonised” land (CAL), or land from which

on land for cash crops) and usually represents a good portion of the household’s income. Using the data represented in the LTBA (cf. Figure 58.1), several indicators can be derived to better understand the system. For instance, x = L&E/HADF could be used to measure the social system’s institutional differentiation; x = Wsub/HAwork can inform on Clemens M. Grünbühel

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the level of subsistence; and x = W − land/W$ gives an indication on the land dependency for income. Similarly, x = EOAL/TAL could be a proxy for the potential for ecosystem services, and x = LIPsub/LIP an indicator for food security. Depending on the objective of the analysis, various other indicators can be derived to characterise the system.

Current applications

Grünbühel and Schandl (2005) have used LTBA, following Giampietro, to analyse national policy on the eradication of shifting cultivation in Lao PDR. This is an example of the multilevel application of LTBA, in that they conducted the analysis for the local level (semi-subsistent farmer community) as well as the national level (using national statistical time and land use accounts). The analysis shows, at both levels, a society well ingrained in subsistence agriculture with minimal disposable income and hardly any capital to invest in the modernisation of the economy. As policies discourage farmers to continue with shifting cultivation, it becomes clear that they cannot easily jump to another mode of production as capital is not sufficiently available to buffer against the risks of transformation. Serrano-Tovar and Giampietro (2014) employ LTBA as part of the larger MuSIASEM approach to identify different farming system types in a multi-scalar manner, which allows for an integrated understanding of possible development options, including the biophysical ramifications of each option across various levels (household, community, region, nation). The approach links social decisions with metabolic and land use outcomes and demonstrates how LTBA is integrated in societal metabolism methodologies across scales, as well as its application to sustainability analysis. Arizpe et al. (2014) apply LTBA to two communities in the Chaco region of Argentina and investigate how the chosen communities respond to the expansion of soy production in the region. They demonstrate how agricultural modernisation alienates cultivators from their land and disassociates time investments from land use. The case studies also show how capital influx and market integration lead to dependency on the formal monetary economy and how both work and land become commodified. This coincides Clemens M. Grünbühel

with the decrease of land resource conservation and an inevitable decline into poverty, with external actors absorbing the bulk of the profits gained from land use. LTBA has not yet found its way into standard analysis of rural systems. Nevertheless, with data accessible through standard field methods or available statistics, its potential has been proven in various applications. Inclusion in rapid rural analysis or participatory rural appraisal would be straightforward and likely beneficial to understand resource use decisions, segmentation, and livelihood types of local socio-ecological systems. Clemens M. Grünbühel

Notes

1. “Time and space are the main dimensions of practical action and adaptation to the natural and social environment, since any use of resources and people in human projects entails arranging activities in space and time so that the right inputs are combined at the right times and places” (Carlstein 1982: 38) 2. However, leisure time becomes more democratised in modern societies with standardised work weeks, mandatory leave, and workforce regulations for all employees. Working hours per capita have decreased due to policies over the last 80 years. 3. Another way capital can be used is to expand the land-time budget by hiring workers or purchasing access to land. 4. The assumption is that the physiological overhead, while variable individually, may not be very elastic (i.e. non-adaptive) across the group. There is much larger flexibility, however, when deciding between work time and social time (Mayrhofer-Grünbühel 2004; Grünbühel & Schandl 2005).

References

Arizpe, N., J. Ramos-Martín, M. Giampietro. 2014. An assessment of the metabolic profile implied by agricultural change in two rural communities in the north of Argentina. Environment, Development and Sustainability 16: 903–24. Boserup, E. 2002/1965. The Conditions of Agricultural Growth. The Economics of Agrarian Change under Population Pressure. Earthscan: London. Carlstein, T. 1982. Time Resources, Society and Ecology. Volume 1: Preindustrial Societies. Allen & Unwin: London. Ellen, R. 1996. Introduction. In: R. Ellen & K. Fukui (Eds.), Redefining Nature: Ecology, Culture and Domestication. Routledge: New York, 1–36. Food and Agriculture Organization (FAO). 2012. Sustainability Pathways: Smallholders and Family Farmers. FAO: Rome. http://​ www​ .fao​.org/​fileadmin/​templates/​nr/​sustainability​

Land-time budget analysis  341 _pathways/​docs/​Factsheet​_SMALLHOLDERS​ .pdf Georgescu-Roegen, N. 1971. The Entropy Law and the Economic Process. Harvard University Press: Cambridge, MA. Gershuny, J. 2000. Changing Times. Work and Leisure in Postindustrial Society. Oxford University Press: Oxford. Giampietro, M. 2004. Multi-Scale Integrated Analysis of Agroecosystems. CRC Press: Boca Raton, FL. Gomiero, T., M. Giampietro 2001. Multiple scale integrated analysis of farming systems: the Thuong Lo commune (Vietnamese uplands) case study. Population and Environment 22(3): 315–52. Groh, D. 1992. Anthropologische Dimensionen der Geschichte. Suhrkamp: Frankfurt. Grünbühel, C.M., H. Schandl. 2005. Using land-time-budgets to analyse farming systems and poverty alleviation policies in the Lao PDR. International Journal of Global Environmental Issues 5(3–4): 142–180. Grünbühel, C.M., L.J. Williams. 2016. Risks, resources and reason: understanding smallholder decisions in farming system development interventions in Eastern Indonesia. Journal of Agriculture and Rural Development

in the Tropics and Subtropics 117(2): 295–308. http://​www​.jarts​.info/​index​.php/​jarts/​article/​ view/​2016101851052 Mayrhofer-Grünbühel, C. 2004. Resource Use Systems and Rural Smallholders. An Analysis of Two Lao Communities. Dissertation. Universität Wien: Vienna. Odum, E.P. 1971. Fundamentals of Ecology. Saunders: Philadelphia. Pastore, G., M. Giampietro, L. Ji. 1999. Conventional and Land-Time Budget Analysis of rural villages in Hubei Province, China. Critical Reviews in Plant Sciences 18(3): 331–57. Rappaport, R.A. 1969. Ritual regulation of environmental relations among a New Guinea people. In: A.P. Vayda (Ed.), Environment and Cultural Behaviour. Natural History Press: Garden City, NJ, 181–201. Sahlins, M. 1974. Stone Age Economics. Tavistock: London. Serrano-Tovar, T., M. Giampietro 2014. Multi-scale integrated analysis of rural Laos: studying metabolic patterns of land uses across different levels and scales. Land Use Policy 36: 155–70. Smil, V. 2017. Energy and Civilization. A History. MIT Press: Cambridge, MA.

Clemens M. Grünbühel

59. Languages of valuation Languages of valuation is a term used to denote the vocabulary or idioms that are used to place a value on nature. For example, an individual may hold that a rock is valuable because of the economic profit that would accrue from mining its minerals, or consider the rock valuable because of its sacredness as the resting place of ancient spirits, as Aboriginal people of Australia consider Uluru. In those examples, profit and sacredness comprise two different languages used to express the value of a rock. One could trace a genealogy of the concept of languages of valuation in two intellectual projects at the heart of ecological economics: the desire to overcome the reductionism of using monetary values for understanding environmental preferences, and the conceptualisation of environmental conflicts as ecological distribution conflicts. From the start, ecological economics emphasised that there is a plurality of beliefs about what is of value when it comes to nature (O’Neill and Spash, 2000), and that environmental values are often not commensurable and so cannot be measured in the same unit (Martínez-Alier et al., 1997). Such recognition of value pluralism and incommensurability led to calls for employing multiple means of valuation in the resolution of environmental conflicts and decision-making (Martínez-Alier et al., 1997). It also led to calls for integrating multiple value articulating institutions in decision-making processes, meaning multiple frames that are invoked in the process of expressing values that regulate and influence differently which values come forward and which are excluded (Vatn, 2005). Similarly, the field underlined the relevance of multiple types of foundations for rational action – namely, consequentialist, rights-based, and procedural concerns – in an effort to go beyond the limited and mechanistic Homo economicus model of rationality at the centre of monetary valuation (Zografos and Paavola, 2008). In parallel, environmental conflicts were conceptualised as conflicts about unequal distributions of ‘goods’ and ‘bads’ from environmental change, or else ecological distribution conflicts (Martínez-Alier, 2002). Ecological

economists argued that environmental and health damages from the expansion of the social metabolism of materially abundant economies represent cost shifts (Kapp, 1975; Aguilera-Klink and Alcántara, 1994), which tend to be unequally distributed. Such unequal distributions often occur at the commodity frontiers, those locations, communities, and ecosystems where the quest to reduce production costs, or generate new opportunities for profit, creates contaminating activities. In those situations, communities or environmental justice organisations challenge inequality by expressing how they value nature and their relation to it in ways that cannot be captured or compared to the language of monetary value (Martínez-Alier, 2008). They use languages such as the sacredness or rights of nature, human rights, territorial rights, national sovereignty, social and environmental justice, prior consent, and so on, which cannot be transformed into a price tag (Martínez-Alier, 2002). This makes it practically impossible to internalise externalities and offer monetary compensation to communities for sacrificing certain values, and to avoid conflict or settle it in a fair manner (Temper et al., 2018). And when monetary value is imposed as the single language of valuation to decide ‘optimal’ solutions in ecological distribution conflicts (as in cases where cost–benefit analysis is imposed as the sole policy evaluation instrument), this involves a questionable exercise of ‘procedural power’ (i.e. the power to determine a decision-making bottom line in the face of irreducible complexity; Martínez-Alier, 2002). Languages of valuation have been instrumental for establishing ecological economics’ normative case for value pluralism in environmental decision-making. Through the concept, the discipline was able to sustain a solid case for diversity, inclusion, and plurality, which emphasised ways in which multiple rationalities, values, and ethics are relevant when valuing the environment, engaging, and ‘resolving’, but also understanding environmental conflicts.

Lines of enquiry

Somehow typical of economic analysis, ecological economics has approached languages of valuation from both positive and normative lenses of analysis. The positive approach has

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sought to explain what happens in those cases where monetary and non-monetary languages of valuation clash, whereas the normative strand has sought to determine instruments for resolving the clash through fair and effective environmental decision-making processes. Investigating instances where non-monetary valuation languages challenge monetary valuation is probably the most common feature of published studies that engage with the concept. Comprehensive analyses of numerous ecological distribution conflicts recorded in the Environmental Justice Atlas identify several occurrences of value system contests where the assumption that externalities can be priced is questioned (Temper et al., 2018). Other studies highlight how the use of monetary valuation by the conservation movement may diverge from the environmentalism of the poor which appeals more to non-economic values (Rodríguez-Labajos and Martínez-Alier, 2013); urban community garden initiatives advancing cultural, reparation, and community cohesion in contrast to languages of green consumption and compact cities (Anguelovski and Martínez-Alier, 2014); commercial logging in Africa, where languages of economic growth are challenged by languages premised on livelihood, custom, and sacredness (Veuthy and Gerber, 2011); calls for energy sovereignty and environmental justice by anti-dam resistance movements in India (Del Bene, 2018); prioritisation of ‘ecological balance’ and environmental quality over monetary compensation in gold-mining projects in Turkey (Avcı et al., 2010); the mobilisation of civil and human rights in mining (Urkidi and Walter, 2011) and palm oil and sugarcane plantation (Mingorría Martinez, 2017) conflicts in Latin America; deployment of languages of sovereignty, democracy, commons, and human rights by Latin American grassroots organisations and scholars against utilitarian understandings of natural goods as ‘commodities’ or ‘strategic natural resources’, an ‘eco-territorial’ turn in environmental struggles that disputes the meaning of development and sustainability, prioritising democratisation of decision-making and the right to say ‘no’ (Svampa, 2015). In some of those challenges, an economic language of valuation does not carry the day (Martínez-Alier et al., 2010). Importantly, beyond blocking environmentally damag-

ing projects, non-monetary values can be relatively successful internally within environmental movements, such as in Mexico, where their use by Indigenous groups opened up spaces of political organisation that enabled the creation of resistance networks (Avila-Calero, 2017). In other cases, though, the diversity of plural and multiple valuation languages mobilised in a conflict can make it difficult to establish paths for alliances (Cardoso, 2018). But it is also not uncommon for monetary valuation to impose itself, especially in cases where institutional (states, municipalities, etc.) and corporate actors emphasise the benefits of economic growth and their ‘trickle-down’ effects for compensating environmental losses (Anguelovski and Martínez-Alier, 2014). Moreover, non-monetary valuation languages may be excluded through legal means (Martínez-Alier et al., 2010), like in the case of the Indian Supreme Court’s 2006 decision over the dismantling of the ocean liner Blue Lady, which prioritised the expression of sustainability in terms of monetary benefit at the national scale (Demaria, 2010). And such exclusions are also connected to illegal exercises of power, involving multiple forms of violence (Navas et al., 2018) that include physical abuse and even the killing of environmental activists (Scheidel et al., 2020) who regularly defend livelihoods and ecosystems by reclaiming human, Indigenous, labour, land, women’s, and ecosystem rights. Another pathway of interaction between monetary and non-monetary values involves compromises or combinations of the two instead of clash. For example, rural communities in the Global South use monetary reparation as a language in tandem with the defence of environmental conditions and human and customary rights (Gerber et al., 2009); Indigenous communities, women activists, grassroots organisations, and citizen groups may demand compensation for damages while simultaneously demanding respect for human rights, Indigenous territorial rights, and sacredness (Martínez-Alier et al., 2010; Anguelovski and Martínez-Alier, 2014); and the climate justice movement has supported monetary calculations of the ‘ecological debt’ (Rodríguez-Labajos and Martínez-Alier, 2013). Environmental movements seem to employ the technical language of Western environmentalism for strategic Christos Zografos

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reasons while combining it with arguments about identity and culture (Temper et al., 2018). In some cases, monetary-based policy tools such as payments for environmental services seem to have the potential to integrate diverse languages of valuation (Kallis et al., 2013). Some support that approaches which allow monetary valuation to strengthen the case of other valuation languages is possible and should be pursued, maintaining that in real life it is non-monetary concerns, such as rights, safety, and ethics, that tend to overrule monetary values (Gsottbauer et al., 2015). But other scholars claim that monetary valuation could be acceptable only if it helps improve the environment while bringing more equality, including not suppressing other valuation languages and value-articulating institutions (Kallis et al., 2013). Suggestions about combining monetary and non-monetary languages of valuation coincide with the more normative agenda of deliberative ecological economics, which seeks to establish formal procedures for accommodating value pluralism through deliberation in environmental decision-making processes. From early on, scholars in ecological economics have worked towards finding ways to operationalise incommensurability in environmental values. Munda’s (1995) development of a social multi-criteria analysis that permits operationalising ‘weak comparability’ of environmental values expressed in different units through his NAIADE model stands out for its capacity to integrate expressions of environmental value in multiple units of measurement. Others have adopted non-positivist, mixed-methods and interpretive policy analysis approaches, such as Q methodology, for making normative contributions to environmental policy (Barry and Proops, 1999) and values (Zografos, 2007). Building on that background, deliberative ecological economics has developed two main lines of work. The first combines deliberation with either monetary group-based valuation or non-monetary (e.g. deliberative multi-criteria analysis) decision-making tools with the purpose to integrate multiple languages of valuation and arrive at collective decisions, either through monetising or by keeping with the incommensurability principle (respectively). A second more critical strand investigates political barriers for expressing and taking into account multiple valuation languages in environmental Christos Zografos

decision-making in an effort to specify conditions for inclusive sustainability politics (Zografos and Howarth, 2008). That work establishes that value plurality does not need to diminish to achieve consensus (Lo, 2013), and that social learning through deliberation may even induce decision-makers to consider ecosystems as priceless (Kenter et al., 2011). More critical takes have highlighted how distributional inequalities may combine with formal decision-making principles (e.g. the ‘public interest’; Zografos, 2017), or informal elements of the decision-making process and technocratic planning tools (Zografos and Martínez-Alier, 2009) to encourage consequentialism and exclusion of idioms that emphasise emotion and personal experience. Such situations create conditions in which multiple valuation languages cannot be expressed or negotiated. Deliberative monetary valuation has been criticised for pretending that two models with radically different ontological presuppositions, such as deliberation (with its collectivist outlook) and monetary valuation (with its individualist outlook), can be combined or held in conjunction (Spash, 2008). Deliberative decision-making has also been criticised for ignoring the practical context of power that surrounds and pervades decision-making spaces, and that can privilege certain voices, such as the voice of reason at the expense of, for example, emotional aspects of human experience (Zografos and Howarth, 2010) that are nevertheless crucial for achieving urgent socio-ecological transformations (Nelson, 2013). Partly in response to such criticisms, some work has been reoriented towards looking at the challenges that direct democracy (itself based on deliberative decision-making) faces as a vehicle for facilitating inclusive, radical socio-ecological transformation projects, such as degrowth (Zografos, 2015) and post-development (Zografos, 2019). Meanwhile, other work (Martínez-Alier, 2019) suggests that the coining and mobilising of terms such as biopiracy, sacrifice zones, green deserts, and others allows environmental movements to not only push for alternative social transformation through deploying new vocabularies, but may also contribute to the pluriverse project, which aspires mobilising a process of intellectual, emotional, ethical, and spiritual

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decolonisation of the idea of ‘development as progress’ (Kothari et al., 2019).

Still to come

The normative approach of deliberative ecological economics has taken root with its distinctive research agenda and scholarship (Kenter et al., 2016, 2015). As deliberative ecological economics is treated in another chapter of this book, I here focus on the future of positive analysis of languages of valuation – with my views by no means representing an exhaustive research agenda. The literature presented here shows that positive analyses have established the role of non-monetary languages within ecological distribution conflicts as challengers to monetary valuations, but also at times as their complementary elements in local struggles to maintain control of lives and ecologies in the face of expansions of the social metabolism of Global North economies. Keeping with the ecological economics tenet of value plurality, such approaches rightly advocate for the inclusion of varied modes of expressing environmental value in sustainability decision-making. Yet, and despite its undeniable relevance, inclusion in decision-making can also go hand-in-hand with inclusion within the reach of government action and control in ways that can demobilise political resistance to resource takeover, the expansion of capitalist relations, and racial, patriarchal, and class domination, notably within the current context of developing just transition climate policy (Andreucci and Zografos, 2022). How do different languages of valuation perform within such political environments? Do all non-monetary languages manage to successfully challenge resource takeovers, or do some become prey of political co-optation? Are, for example, non-monetary languages romanticised as pre-modern expressions that disclose the need to improve populations and ecologies still outside the contours of capitalist value creation through modernisation? And under what socio-political and environmental conditions each of those things may happen? Answering such questions would help expand the current focus of analysis on the political work that languages of valuation do. This looks at how certain – and, in particular, non-monetary – environmental values act at the collective level, providing

a common language that allows communities and environmental justice organisations to contest power relations within ecological distribution conflicts; and vice versa, how other, usually utilitarian languages provide political resources for others (e.g. corporations, the state) attempting to sustain power relations. This approach has proven invaluable for visibilising and analysing resistance to environmentally unjust expansions of Global North social metabolisms, particularly through projects such as the Environmental Justice Atlas, which today documents more than 3500 such cases (ejatlas.org/). Yet, clashes over environmental values operate also at a personal, subjective level, which implies that their political work also takes place at those levels. This is because environmental values are not always ‘cognitive’ but are also emotional; they are about what one is allowed to remember, feel, enjoy, or live (Velicu, 2015). Similarly, environmental conflicts are also emotional conflicts (Sultana, 2011) that may reflect resistance to the moulding of certain subjectivities occurring at different levels ranging from the psychological to the geographical and the personal-political (González-Hidalgo and Zografos, 2020). Considering also the pivotal role of emotions in shaping moral action (Nelson, 2013), investigating connections between emotions, languages of valuation, and environmental conflict could shed light upon crucial aspects of ecological distribution conflicts. For example, mappings of emotional (e.g. anger, confusion, desire, distress, joy, fear, sadness, etc.) inclusions and exclusions expressed through diverse languages of valuation could help to better explain successful and less successful resistance in conflicts, and bring to the fore subjectivity elements that contribute to the emergence of local and global movements for environmental justice. This could be pursued with more macro studies as part or akin to the Environmental Justice Atlas project, or with individual and comparative case study research that would contribute to analytical generalisations. With such research, the concept of languages of valuation could be connected to a booming literature on emotions within feminist studies, political ecology, and environmental social science (Nightingale et al., 2021; González-Hidalgo and Zografos, 2017; Christos Zografos

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Sultana, 2015; Singh, 2013; Nightingale, 2012). Until now, research on languages of valuation has transformed what was a debate confined to disagreements between environmentally minded economists about the capacity of money to ‘capture’ ‘real’ preferences and design optimal environmental policy, to a broad and diverse scholarship which connects discussions between environmental policy, social movements, sustainability governance, environmental philosophy and ethics, institutional economics, environmental history, and political ecology-minded scholars. Future research has potential to expand and nuance even more such connections, and establish even further languages of valuation as a central analytical concept for understanding environmental conflict and governance. Christos Zografos

Demaria, F., 2010. Shipbreaking at Alang–Sosiya (India): An ecological distribution conflict. Ecological Economics, Special Section: Ecological Distribution Conflicts 70, 250–60. https://​doi​.org/​10​.1016/​j​.ecolecon​.2010​.09​.006 Gerber, J.-F., Veuthey, S., Martínez-Alier, J., 2009. Linking political ecology with ecological economics in tree plantation conflicts in Cameroon and Ecuador. Ecological Economics 68, 2885–9. https://​doi​.org/​10​.1016/​j​.ecolecon​ .2009​.06​.029 González-Hidalgo, M., Zografos, C., 2017. How sovereignty claims and ‘negative’ emotions influence the process of subject-making: Evidence from a case of conflict over tree plantations from Southern Chile. Geoforum 78, 61–73. https://​doi​.org/​10​.1016/​j​.geoforum​ .2016​.11​.012 González-Hidalgo, M., Zografos, C., 2020. Emotions, power, and environmental conflict: Expanding the ‘emotional turn’ in political ecology. Progress in Human Geography 44, 235–55. https://​doi​.org/​10​.1177/​ 0309132518824644 Gsottbauer, E., Logar, I., van den Bergh, J., 2015. References Towards a fair, constructive and consistent Aguilera-Klink, F., Alcántara, V., 1994. De la criticism of all valuation languages: Comment economía ambiental a la economía ecológica. on Kallis et al. (2013). Ecological Economics Universidad Privada del Norte, Barcelona. 112, 164–9. https://​doi​.org/​10​.1016/​j​.ecolecon​ Andreucci, D., Zografos, C., 2022. Between .2014​.12​.014 improvement and sacrifice: Othering and the Kallis, G., Gómez-Baggethun, E., Zografos, C., (bio) political ecology of climate change. 2013. To value or not to value? That is not the Political Geography 92, 102512. question. Ecological Economics 94, 97–105. Anguelovski, I., Martínez-Alier, J., 2014. The https://​doi​.org/​10​.1016/​j​.ecolecon​.2013​.07​.002 ‘Environmentalism of the Poor’ revisited: Kapp, K.W., 1975. The Social Costs of Private Territory and place in disconnected global Enterprise, 2nd ed. Schocken Books. struggles. Ecological Economics 102, 167–76. Kenter, J.O., Bryce, R., Christie, M., Cooper, N., https://​doi​.org/​10​.1016/​j​.ecolecon​.2014​.04​.005 Hockley, N., Irvine, K.N., Fazey, I., O’Brien, Avcı, D., Adaman, F., Özkaynak, B., 2010. L., Orchard-Webb, J., Ravenscroft, N., 2016. Valuation languages in environmental conShared values and deliberative valuation: Future flicts: How stakeholders oppose or support directions. Ecosystem Services 21, 358–71. gold mining at Mount Ida, Turkey. Ecological Economics, Special Section: Ecological Kenter, J.O., Hyde, T., Christie, M., Fazey, I., 2011. The importance of deliberation in valuing Distribution Conflicts 70, 228–38. https://​doi​ ecosystem services in developing countries— .org/​10​.1016/​j​.ecolecon​.2010​.05​.009 Evidence from the Solomon Islands. Global Avila-Calero, S., 2017. Contesting energy transiEnvironmental Change 21, 505–21. https://​doi​ tions: Wind power and conflicts in the Isthmus .org/​10​.1016/​j​.gloenvcha​.2011​.01​.001 of Tehuantepec. Journal of Political Ecology 24, 992–1012. https://​doi​.org/​10​.2458/​v24i1​ Kenter, J.O., O’Brien, L., Hockley, N., Ravenscroft, N., Fazey, I., Irvine, K.N., Reed, .20979 M.S., Christie, M., Brady, E., Bryce, R., 2015. Barry, J., Proops, J., 1999. Seeking sustainabilWhat are shared and social values of ecosysity discourses with Q methodology. Ecological tems? Ecological Economics 111, 86–99. Economics 28, 337–45. Cardoso, A., 2018. Valuation languages along the Kothari, A., Salleh, A., Escobar, A., Demaria, F. and Acosta, A. (Eds.), 2019. Pluriverse: coal chain from Colombia to the Netherlands A Post-Development Dictionary. Tulika Books and to Turkey. Ecological Economics 146, and Authorsupfront. 44–59. https://​doi​.org/​10​.1016/​j​.ecolecon​.2017​ Lo, A.Y., 2013. Agreeing to pay under value disa.09​.012 greement: Reconceptualizing preference transDel Bene, D., 2018. Hydropower and ecologiformation in terms of pluralism with evidence cal conflicts. From resistance to transformafrom small-group deliberations on climate tions (PhD thesis). Universitat Autònoma de change. Ecological Economics 87, 84–94. Barcelona. https://​doi​.org/​10​.1016/​j​.ecolecon​.2012​.12​.014

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.gloenvcha​.2020​.102104 Singh, N.M., 2013. The affective labor of growing forests and the becoming of environmental subjects: Rethinking environmentality in Odisha, India. Geoforum 47, 189–98. https://​doi​.org/​10​ .1016/​j​.geoforum​.2013​.01​.010 Spash, C.L., 2008. Deliberative monetary valuation: Literature, limitations and perspectives, in: Deliberative Ecological Economics, Zografos, C., and Howarth, R.B. (Eds.), Oxford University Press, pp. 21–49. Sultana, F., 2011. Suffering for water, suffering from water: Emotional geographies of resource access, control and conflict. Geoforum, Themed Issue: New Feminist Political Ecologies 42, 163–72. https://​doi​.org/​10​.1016/​j​.geoforum​ .2010​.12​.002 Sultana, F., 2015. Emotional political ecology, in: The International Handbook of Political Ecology, in: Bryant, R.L. (Ed.), Edward Elgar Publishing, pp. 633–45. Svampa, M., 2015. The ‘commodities consensus’ and valuation languages in Latin America. Alternautas 2, 45–59. Temper, L., Demaria, F., Scheidel, A., Del Bene, D., Martínez-Alier, J., 2018. The Global Environmental Justice Atlas (EJAtlas): Ecological distribution conflicts as forces for sustainability. Sustainability Science 13, 573–84. https://​doi​.org/​10​.1007/​s11625​-018​ -0563​-4 Urkidi, L., Walter, M., 2011. Dimensions of environmental justice in anti-gold mining movements in Latin America. Geoforum 42, 683–95. https://​doi​.org/​10​.1016/​j​.geoforum​.2011​.06​ .003 Vatn, A., 2005. Institutions and the Environment. Edward Elgar Publishing. Velicu, I., 2015. Demonizing the sensible and the ‘revolution of our generation’ in Rosia, Montana. Globalizations 12, 846–58. https://​doi​ .org/​10​.1080/​14747731​.2015​.1100858 Veuthy, S., Gerber, J.-F., 2011. Valuation contests over the commoditisation of the Moabi Tree in south-eastern Cameroon. Environmental Values 20, 239–64. https://​doi​.org/​10​.3197/​ 0963​27111X1299​7574391805 Zografos, C., 2019. Direct democracy, in: Pluriverse: A Post-Development Dictionary, Kothari, A., Salleh, A., Escobar, A., Demaria, F., and Acosta, A. (Eds.), Tulika Books and Authorsupfront, pp. 154–7. Zografos, C., 2017. Flows of sediment, flows of insecurity: Climate change adaptation and the social contract in the Ebro Delta, Catalonia. Geoforum 80, 49–60. https://​doi​.org/​10​.1016/​j​ .geoforum​.2017​.01​.004 Zografos, C., 2015. Démocratie directe [Direct democracy], in: D’Alisa, G., Demaria, F., and Kallis, G. (Eds.), Décroissance. Vocabulaire pour une nouvelle ère [Degrowth. Vocabulary

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348  Elgar encyclopedia of ecological economics for a New Era], Erudit, pp. 187–94. Zografos, C., 2007. Rurality discourses and the role of the social enterprise in regenerating rural Scotland. Journal of Rural Studies 23, 38–51. Zografos, C., Howarth, R.B., 2008. Towards a deliberative ecological economics, in: Deliberative Ecological Economics, Zografos, C., and Howarth, R.B. (Eds.), Oxford University Press, pp. 1–20. Zografos, C., Howarth, R.B., 2010. Deliberative ecological economics for sustainability govern-

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ance. Sustainability 2, 3399–3417. Zografos, C., Martínez-Alier, J., 2009. The politics of landscape value: A case study of wind farm conflict in rural Catalonia. Environment and Planning A 41, 1726–44. Zografos, C., Paavola, J., 2008. Critical perspectives on human action and deliberative ecological economics, in: Deliberative Ecological Economics, Zografos, C., and Howarth, R.B. (Eds.), Oxford University Press, pp. 146–66.

60. The laws of thermodynamics There are four Laws of Thermodynamics. They are called the Zeroth, First, Second, and Third Laws because the Zeroth Law was developed last but is now seen to be a fundamental underpinning of the other laws. The Second Law has three equivalent statements, and the Third Law has two equivalent statements; proof of these equivalences can be found in Zemansky (1968). The term “entropy” is explained below. ● Zeroth Law: Two systems which are in thermal equilibrium with a third are in thermal equilibrium with each other (Dugdale, 1996, p. 13). ● First Law: During a process in which no heat is exchanged with the environment, the work done is only a function of the initial and final states of the system, not of the path. Furthermore, during any process, the change in the initial energy of the system, ​U ​ f​ ​ − ​Ui​ ​​, is equal to the heat flow into the system, ​Q​, minus the net work done by the system,​ W :    ​Uf​ ​ − ​Ui​ ​  =  Q − W​(Dugdale, 1996, p. 20; Zemansky, 1968, pp. 78–9). ● Second Law, Kelvin–Planck statement: No process is possible whose sole result is the absorption of heat from a reservoir and the conversion of this heat into work (Zemansky, 1968, p. 178). ● Second Law, Clausius statement: No process is possible whose sole result is the transfer of heat from a cooler to a hotter body (Zemansky, 1968, p. 184). ● Second Law, entropy statement: In an isolated system, entropy is nondecreasing (Zemansky, 1968, p. 234). ● Third Law, Unattainability statement: It is impossible to reach absolute zero by any finite number of processes (Zemansky, 1968, p. 498; Dugdale, 1996, p. 177). ● Third Law, Nerst–Simon statement: In the limit as temperature goes to 0 K, the entropy change of any reaction is zero (Zemansky, 1968, p. 498; Rao, 1985, p. 257; Dugdale, 1996, pp. 160–61). In the Zeroth Law, systems are in “thermal equilibrium” with each other if heat could flow between them but does not, as evidenced

by the observation that no characteristics of the systems change over time. The Zeroth Law is needed to be able to conceive of the idea of temperature, which is the thing that systems in thermal equilibrium with each other have in common. The First Law reflects conservation of energy, and clarifies that heat and work are measured in the same units as each other—the joule, denoted “J,” the metric unit of energy (the English units being calories or British thermal units)—and absorption of one unit of heat by a system increases its internal energy in the same way as when one unit of work is done on the system. “Work” is equal to force times distance, and since it involves distance, it is a macroscopic phenomenon; heat is a microscopic phenomenon. Internal energy changes are defined, but the internal energy level has an arbitrary origin (just like gravitational or magnetic potential energy). Heat is often referred to as a “flow” even though its unit is joules, not joules per second (a “watt”), because both heat and work happen over time to cause changes in the stock of U ​ ​. The Third Law concerns behavior near absolute zero, which has little importance for ecological economics except for a few points noted below.

Entropy in classical thermodynamics

Let T ​​ stand for absolute temperature, measured in Kelvin, denoted K (prior to the mid-1960s, it was measured in “degrees Kelvin” and denoted °K). Kelvin temperature is 273.15° higher than Celsius temperature, and is named after William Thomson, also known as Lord Kelvin. Let ​Q​ stand for the flow of heat into a material (or system, which is a collection of materials). In “Classical Thermodynamics” developed in the mid-19th century, the change in the entropy S of the material (or system) is defined to be: d ​Q​  ​

​dS  ​=  _      ​  ​  Trev 

(​ 60.1)​​

where the subscript “rev” stands for a reversible process; physicists call a process “reversible” if it involves no dissipative effects, such as friction, viscosity, inelasticity, electrical resistance, or magnetic hysteresis (Zemansky, 1968, pp. 193, 215; Mackowiak, 1965, p. 59). As Zemansky adds (1968, p. 225), if a system undergoes an irreversible process between an

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initial equilibrium state i and a final equilibrium state f, the entropy change of the system is equal to the integral from i to f of ​dQ / T​ taken over any reversible path from i to f. No integral is taken over the original irreversible path. Suppose one has an isolated system containing two bodies, one hot and one cold, placed in thermal contact. When a given quantity of heat, Q0, flows from the hotter body to the colder one, the change in entropy of the system is: ​−  Q ​ ​ ​

​+  Q ​ ​ ​

hot

cold 

​ΔS  ​=  _ ​ ​T​  0 ​​  ​+  _ ​ ​T​  0 ​ ​

Since ​Thot  ​  ​  >  ​Tcold  ​  ​,  ΔS  >  0​. If heat were to flow in the opposite direction, away from the colder body and toward the warmer body, then the Q0 terms in the above equation would change sign, and ​ΔS​ would be negative, violating the Second Law. Heat can flow from a colder body to a warmer one, and thus, entropy can decrease, in part of a system, or in a system that is not isolated, but not in an isolated system. The statement “in an isolated system, entropy never decreases” implicitly means that “entropy never decreases as time goes forward.” It follows that the Second Law is connected with, manifests, or perhaps even provides “Time’s Arrow”—that is, the direction in which time flows. Earlier physical laws, such as Newton’s Laws of Motion, are symmetric in time: in their mathematical forms, substituting −t for time t generates true statements, which is called “T-symmetry.” The Second Law is time-asymmetric. Given two dates, ​t​1​ ​   and   ​t2​ ​  >  t​1​ ​​, the entropy of an isolated system at t 1 is less than the entropy of the system at ​t2​ ​ :   ​S1​ ​