December 11, 2017 | Author: aaguilard | Category: Biodiversity, Ecosystem Services, Forests, Grassland, Conservation Biology
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Biodiversity Conservation in Latin America and the Caribbean

Latin America and the Caribbean (LAC) region is exceptionally biodiverse. It contains about half of the world’s remaining tropical forests, nearly one-fifth of its coastal habitats, and some of its most productive agricultural and marine areas. But agriculture, fishing, and other human activities linked to rapid population and economic growth increasingly threaten that biodiversity. Moreover, poverty, weak regulatory capacity, and limited political will hamper conservation. Given this dilemma, it is critically important to design conservation strategies on the basis of the best available information about both biodiversity and the track records of the various policies that have been used to protect it. This rigorously researched book describes the status of biodiversity in LAC, the main threats to this biodiversity, and the drivers of these threats. It identifies the main policies being used to conserve biodiversity and assesses their effectiveness and potential for further implementation. It proposes five specific lines of practical action for conserving LAC biodiversity, based on: green agriculture; strengthening terrestrial protected areas and co-management; improving environmental governance; strengthening coastal and marine resource management; and improving biodiversity data and policy evaluation. Allen Blackman is Thomas Klutznick Senior Fellow at Resources for the Future, USA. Rebecca Epanchin-Niell is Fellow at Resources for the Future, USA. Juha Siikamäki is Associate Research Director and Fellow at Resources for the Future, USA. Daniel Velez-Lopez is a PhD student in Public Policy at the Harvard Kennedy School of Government and former Research Assistant at Resources for the Future, USA.

About Resources for the Future and RFF Press

Resources for the Future (RFF) improves environmental and natural resource policymaking worldwide through independent social science research of the highest caliber. Founded in 1952, RFF pioneered the application of economics as a tool for developing more effective policy about the use and conservation of natural resources. Its scholars continue to employ social science methods to analyze critical issues concerning pollution control, energy policy, land and water use, hazardous waste, climate change, biodiversity, and the environmental challenges of developing countries. RFF Press supports the mission of RFF by publishing book-length works that present a broad range of approaches to the study of natural resources and the environment. Its authors and editors include RFF staff, researchers from the larger academic and policy communities, and journalists. Audiences for publications by RFF Press include all of the participants in the policymaking process—scholars, the media, advocacy groups, NGOs, professionals in business and government, and the public.

Resources for the Future Board of Directors Board Leadership W. Bowman Cutter Chair John M. Deutch Vice Chair Frank E. Loy Vice Chair Lawrence H. Linden Treasurer Philip R. Sharp President

Board Members Vicky A. Bailey Anthony Bernhardt Trudy Ann Cameron Red Cavaney Mohamed T. El-Ashry Linda J. Fisher C. Boyden Gray

David Hawkins Rick Holley Peter R. Kagan Sally Katzen Rubén Kraiem Bob Litterman Richard G. Newell Henry Schacht Richard Schmalensee Lisa A. Stewart Joseph Stiglitz Mark R. Tercek

Chair Emeriti Darius W. Gaskins, Jr. Robert E. Grady

Editorial Advisers for RFF Press Walter A. Rosenbaum, University of Florida Jeffrey K. Stine, Smithsonian Institution

“At last: the handbook on biodiversity conservation in Latin America and the Caribbean we all have needed... with all the considerations necessary for best practice choices... a revolutionary contribution.” – Tom Lovejoy, University Professor, George Mason University and Senior Fellow, United Nations Foundation. “A great addition to literature, this book starts by describing LAC biodiversity’s status and progresses to a critical study of the main conservation policies. It is here that the book excels becoming a fascinating read for those involved in the field and a compulsory one from the management and education perspective.” – Francisco Alpízar, Founder, Latin American and Caribbean Environmental Economics Program (LACEEP), Director, Economics and Environment for Development (EfD-CATIE) and Associate Professor, Department of Economics, University of Gothenburg. “This book has been instrumental in setting new directions for conservation investments at the Interamerican Development Bank and provides the foundation for more effective policy in the future.” – Michele Lemay, Natural Resources Lead Specialist, Inter-American Development Bank. “This book provides a wealth of data and information, a clear-eyed assessment of the challenges to biodiversity conservation in the region, and a valuable framework for prioritizing policies. It makes it clear that mainstreaming biodiversity will require a continuous and coherent process in which early and well planned commitments will reduce overall costs.” – Carlos Manuel Rodríguez, Vice President, Conservation International and Former Minister of Environment, Costa Rica.

Biodiversity Conservation in Latin America and the Caribbean Prioritizing policies Allen Blackman, Rebecca Epanchin-Niell, Juha Siikamäki, and Daniel Velez-Lopez

First published 2014 by RFF Press Taylor & Francis, 2 Park Square, Milton Park, Abingdon, Oxon OX14 4RN and by RFF Press Routledge, 711 Third Avenue, New York, NY 10017 RFF Press is an imprint of the Taylor & Francis Group, an informa business © 2014 Allen Blackman, Rebecca Epanchin-Niell, Juha Siikamäki, and Daniel Velez-Lopez The right of Allen Blackman, Rebecca Epanchin-Niell, Juha Siikamäki, and Daniel Velez-Lopez to be identified as authors of this work has been asserted by them in accordance with sections 77 and 78 of the Copyright, Designs and Patents Act 1988. All rights reserved. No part of this book may be reprinted or reproduced or utilized in any form or by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying and recording, or in any information storage or retrieval system, without permission in writing from the publishers. Trademark notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation without intent to infringe. British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging-in-Publication Data Biodiversity conservation in Latin America and the Caribbean : prioritizing policies / Allen Blackman, Rebecca Epanchin-Niell, Juha Siikamaki, [editors]. pages cm. – (Environment for development) 1. Biodiversity conservation–Latin America. 2. Biodiversity conservation–Caribbean Area. 3. Environmental policy–Latin America. 4. Environmental policy–Caribbean Area. 5. Latin America–Environmental conditions. 6. Caribbean Area–Environmental conditions. I. Blackman, Allen. QH77.L25B59 2014 333.95’1609729–dc23 2013042508 ISBN: 978-0-415-73096-9 (hbk) ISBN: 978-1-315-84843-3 (ebk) Typeset in Times New Roman by Cenveo Publisher Services

Abbreviated contents


Introduction 1.1 1.2 1.3 1.4



Background 1 Objectives 2 Methods 2 Organization 3

Status and trends


2.1 Terrestrial and freshwater systems 4 2.2 Coastal and marine systems 18 3

Policies A 3.1 3.2 3.3 3.4 3.5 3.6

Regulatory and comanagement 27 Terrestrial protected areas 27 Forest comanagement 33 Land-use planning 36 Fisheries management 39 Wastewater treatment 43 Environmental governance 45

B Market-based approaches 49 3.7 Subsidy reform 49 3.8 Payments for environmental services 54 3.9 Eco-certification 58 3.10 Ecotourism 59 3.11 Bioprospecting 61 3.12 Mitigation offsets and banking 62 C 3.13 3.14 3.15 3.16 3.17

Other 66 National environmental accounting 66 Corporate social responsibility 68 Greening agriculture 72 Targeting, data, and evaluation 75 Reduced emissions from deforestation and degredation 77


viii 4

Abbreviated contents Lines of action 4.1 4.2 4.3 4.4 4.5 4.6



Green agriculture 86 Strengthen terrestrial protected areas and comanagement 89 Improve environmental governance 91 Strengthen coastal and marine resource management 94 Improve biodiversity data and policy evaluation 96 Policies omitted from lines of action 98

Latin America and Caribbean biodiversity actors


Appendices References Index

101 130 154


List of illustrations About the authors Acknowledgments 1

Introduction 1.1 1.2 1.3 1.4


xiv xvi xvii 1

Background 1 Objectives 2 Methods 2 Organization 3

Status and trends 2.1 Terrestrial and freshwater systems 4 2.1.1 Biophysical environment 4 2.1.2 Status and trends 3 Species richness and overall threat levels 5 Terrestrial ecosystems 6 Freshwater ecosystems 8 2.1.3 Threats 9 IUCN Red List data 9 Habitat loss and degradation 10 Invasive species 15 Climate change 16 Overexploitation 17 Pollution 18 2.2 Coastal and marine systems 18 2.2.1 Biophysical environment 18 2.2.2 Status and trends 19 Species richness 19 Coastal ecosystems 20 2.2.3 Threats 21 Habitat loss and degradation 22 Pollution 22



Contents Overexploitation 23 Climate change 24 Invasive species 25


Policies A Regulatory and comanagement 27 3.1 Terrestrial protected areas 27 3.1.1 Description 27 3.1.2 Status and trends 27 Protected area coverage standards 27 Coverage by region 28 Coverage by country 28 Coverage by type of protection 28 Coverage by biome 29 3.1.3 Issues 29 Coverage gaps 29 Size and fragmentation 30 Management 30 Financial resources 30 Local communities 31 3.1.4 Effectiveness 32 3.1.5 Easements 33 3.2 Forest comanagement 33 3.2.1 Description 33 3.2.2 Status and trends 34 3.2.3 Issues 34 Stemming forest loss 34 Spurring forest loss 35 3.2.4 Effectiveness 35 3.3 Land-use planning 36 3.3.1 Description 36 3.3.2 Status and trends 37 3.3.3 Issues 38 3.4 Fisheries management 39 3.4.1 Description 39 3.4.2 Status and trends 39 Economic importance 39 Fleet and market characteristics 39 Production in capture fisheries 40 3.4.3 Issues 40 3.4.4 Comanagement 42


Contents 3.5 Wastewater treatment 43 3.5.1 Description 43 3.5.2 Status and trends 43 3.5.3 Issues 44 3.6 Environmental governance 45 3.6.1 Description 45 3.6.2 Status and trends 45 Environmental laws, policies, and programs 45 Governance effectiveness 45 Environmental capacity and mainstreaming 47 3.6.3 Issues 48 B Market-based approaches 49 3.7 Subsidy reform 49 3.7.1 Description 49 3.7.2 Status and trends 49 Agriculture 49 Water 51 Fishing 51 Energy 52 3.7.3 Issues 52 3.7.4 Evidence 53 3.8 Payments for environmental services 54 3.8.1 Description 54 3.8.2 Status and trends 54 Overall level of implementation 54 Level of implementation by subregion 55 3.8.3 Issues 55 The case for payments for ecosystem services 55 Barriers to effectiveness 55 3.8.4 Evidence 57 3.9 Eco-certification 58 3.9.1 Description 58 3.9.2 Status and trends 58 3.9.3 Issues 58 3.9.4 Evidence 59 3.10 Ecotourism 59 3.10.1 Description 59 3.10.2 Status and trends 60 3.10.3 Issues and evidence 60 3.11 Bioprospecting 61 3.11.1 Description 61 3.11.2 Status and trends 61



Contents 3.11.3 Issues 61 3.11.4 Evidence 62 3.12 Mitigation offsets and banking 62 3.12.1 Description 62 3.12.2 Status and trends 63 Mitigation offsets 63 Mitigation banking 63 Tradable development rights 64 3.12.3 Issues 65 C Other 66 3.13 National environmental accounting 66 3.13.1 Description 66 3.13.2 Status and trends 66 3.13.3 Issues 67 3.13.4 Evidence 68 3.14 Corporate social responsibility 68 3.14.1 Description 68 3.14.2 Status and trends 68 Overall levels of CSR, by type of participant 68 International pressures 69 3.14.3 Issues 70 The case for CSR 70 Barriers to effectiveness 71 3.14.4 Evidence 71 3.15 Greening agriculture 72 3.15.1 Description 72 3.15.2 Status and trends 73 3.15.3 Issues 73 Preventing extensification 73 Reducing adverse effects of intensification 74 Biofuels 74 3.15.4 Effectiveness 75 3.16 Targeting, data, and evaluation 75 3.16.1 Description 75 3.16.2 Status and trends 76 3.16.3 Issues 76 3.17 REDD 77 3.17.1 Description 77 3.17.2 Issues 78 Scope 78 Reference levels 78 Monitoring 79 Leakage 79



3.17.4 4

xiii Permanence 79 Capacity 79 Co-benefits and safeguards 80 Financing 80 Status and trends 80 Progress under UNFCC 80 Progress outside of the UNFCCC 81 Effectiveness 83

Lines of action


4.1 Green agriculture 86 4.1.1 Rationale 86 4.1.2 Recommendations 86 4.1.3 Expected benefits and indicators 87 4.2 Strengthen terrestrial protected areas and comanagement 89 4.2.1 Rationale 89 4.2.2 Recommendations 90 4.2.3 Expected benefits and indicators 90 4.3 Improve environmental governance 91 4.3.1 Rationale 91 4.3.2 Recommendations 92 4.3.3 Expected benefits and indicators 92 4.4 Strengthen coastal and marine resource management 94 4.4.1 Rationale 94 4.4.2 Recommendations 94 4.4.3 Expected benefits and indicators 95 4.5 Improve biodiversity data and policy evaluation 96 4.5.1 Rationale 96 4.5.2 Recommendations 97 4.5.3 Expected benefits and indicators 97 4.6 Policies omitted from lines of action 98 5

Latin America and Caribbean biodiversity actors


Appendix 1: stakeholder interviews Appendix 2: country-level data Appendix 3: LAC biodiversity actors References Index

101 103 125 130 154

List of illustrations

Tables 2.1-1 2.1-2 2.1-3 2.1-4

2.1-5 2.1-6 2.1-7 3.1-1 3.1-2 3.1-3 3.1-4 3.2-1 3.2-2 3.4-1 3.5-1 3.6-1 3.6-2 3.6-3 3.7-1 3.7-2 3.7-3 3.7-4 3.7-5 3.8-1 3.13-1

Forest area in LAC, by type and use, 2010 Trends in forest area in LAC, 1990–2010 Trends in landcover by biome/region in LAC, 2001–2010 Percentage of IUCN Red List species (vulnerable – extinct) in LAC identifying each threat category as “primary,” by ecosystem type Area devoted to agricultural crops and pasture, by year and subregion Market value of agricultural output, by year and subregion LAC soybean and sugar cane production, by year and subregion Percentage of terrestrial and marine area protected, by region Percentage of terrestrial protected areas, by type and region, 2000 Percentage of terrestrial protected areas, by biome, region, and year Protected area management costs and financial gaps in selected LAC countries, 2010 Forest tenure by subregion, 2005 Forest tenure by subregion for 39 countries with most tropical forest, 2008 Challenges and solutions to unsustainability of fisheries Percentage of sewer water with some type of treatment, by country, 2000 National forest policies, programs, and laws in LAC, by subregion Worldwide Governance Index indicators for LAC (–2.5 to +2.5), 2010 Human resource levels in LAC, 2008 Nominal rate of assistance (NRA) in eight LAC countries, 1965–2004 Gross subsidy equivalents of assistance to farmers in eight LAC countries, 1965–2004 Per person gross subsidy equivalents of assistance to farmers in eight LAC countries, 1965–2004 Spending on irrigation in 13 LAC countries, 2001 Pre-tax energy subsidies as a percentage of GDP PES schemes, by region LAC countries with environmental accounting programs, 2012

6 7 7

10 11 12 12 28 29 29 31 34 35 42 44 45 46 47 50 50 50 51 53 54 66

List of illustrations 3.14-1 3.15-1 3.17-1 3.17-2 4.1 A2.1-1 A2.1-2 A2.1-5 A2.1-6 A2.1-7 A2.2-S1 A2.2-S2 A2.2-S3 A3.1-1 A3.1-S1 A3.2-1 A3.2-S1 A3.4-S1 A3.4-S2 A3.6-1 A3.8-S1 A3.8-S2 A3.15-1

Corporate social responsibility in selected countries in the Americas, by type of participation, 2003 Fertilizer use and intensity, by subregion, 1990–2008 LAC countries participating in major multilateral REDD readiness schemes Total REDD financing reported by funders, by country, 2013 Principal criteria met by each line of action Forest area in LAC, by country, type and use, 2010 Trends in forest area, by country, 1990–2010 Area devoted to agricultural crops and pasture, by country and year Market value of agricultural output, by country and year LAC soybean and sugar cane production, by country and year Population within 100 km of the coastline, by country Current mangrove area and loss rates, by country Landings from coastal fisheries, by country, 2008 Terrestrial and marine area protected, by country and year Government contribution to PAs, by country, 2007 Forest ownership and management rights, by country, 2005 Forest area under community and indigenous tenure regimes, by country, 2012 Fisheries contribution to GDP, by country, 2008 Fisheries employment, by country, 2008 National forest policies, programs, and laws in LAC, by country Watershed PES programs in South America Watershed PES programs in Mesoamerica and the Caribbean Fertilizer use and intensity, by country and year

xv 69 73 82 83 86 104 106 108 109 110 111 112 113 114 115 116 118 118 119 120 122 123 124

Figures 2.2-1 2.2-2

Landings from coastal fisheries in LAC Landings from coastal fisheries in the Caribbean

Plates 2.1-1 2.1-2 2.1-3 2.1-4 2.2-1 2.2-2 2.2-3 2.2-4

Terrestrial biomes of Latin America and the Caribbean Status of LAC ecoregions Global biodiversity hotspots of Latin America and the Caribbean Surface water abstraction stress in LAC Coastal biogeographic provinces in LAC Insular Caribbean and Mesoamerican biodiversity hotspots Mangrove distribution in Americas Fishing densities in coastal LAC

24 25

About the authors

Allen Blackman is Thomas Klutznick Senior Fellow at Resources for the Future. He holds a PhD in Economics from the University of Texas, Austin and a BA in International Relations from the University of Pennsylvania. His research focuses on environmental and natural resource policy in Latin America. Rebecca Epanchin-Niell is Fellow at Resources for the Future. She received a PhD in Agricultural and Resource Economics from the University of California, Davis, MS degrees in Biology and Applied Economics from University of Nevada, Reno, and a BS from Stanford University. Her research tackles issues at the intersection of ecology and economics. Juha Siikamäki is Associate Research Director and Fellow at Resources for the Future. He has a PhD from the University of California, Davis in Environmental Policy Analysis. His research focuses on economic analyses of ecosystem services and biodiversity, especially economic valuation and conservation prioritization. Daniel Velez-Lopez is a PhD student in Public Policy at the Harvard Kennedy School of Government and former Research Assistant at Resources for the Future. He has a BA in Economics and Mathematics from the University of Maryland. His research focuses on environmental policy and political economy in developing countries.


This book evolved from a 2012 report commissioned by the Inter American Development Bank (IDB) to inform the design of its Biodiversity and Ecosystem Services Program. We are grateful to Michele Lemay, Rebecca Benner, Carolina Jaramillo, Erivelthon Lima, and Paul Winters at IDB and to Tim Hardwick and Ashley Wright at RFF Press for multifaceted assistance; our interviewees (listed in Appendix 1) for valuable input; two anonymous reviewers and participants in an April 2012 IDB peer review workshop for helpful comments and suggestions; and Anne Riddle, Jessica Chu, and Sally Atwater for assistance with graphics and editing.

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1.1 Background Although the term biological diversity is often used narrowly to refer to the number of species in a defined geographical area, it has a much broader meaning. It denotes the diversity of living things at all levels, ranging from genetic material in individual organisms to continental ecosystems like forests and grasslands. In both the broad and narrow senses, the region comprising Latin America and the Caribbean (LAC) is exceptionally biodiverse. Stretching from the Northern Hemisphere through tropical, subtropical, and temperate regions to the icy waters off Antarctica, it comprises areas with very different topography and climate that host a variety of ecosystems. It contains about half of the world’s remaining tropical forests, nearly one-fifth of global coastal habitats, five of the world’s 20 longest rivers, and some of the most productive agricultural and marine areas on Earth. And it supports a staggering array of plants and animals. About one-third of all known mammals and an even higher share of reptiles, birds, and amphibians are found there, many of which are endemic – that is, found nowhere else. LAC’s biodiversity has incalculable value. Part is intrinsic: it does not depend on human use. In addition, LAC biodiversity directly underpins a broad range of human activities – such as agriculture, aquaculture, and nature tourism – that generate goods and services sold in markets. These activities provide food, income, and employment to the people in the region. LAC biodiversity also generates a wide array of ecosystem services that are not (normally) bought and sold in markets, including water purification, oxygen creation, maintenance of soil productivity, waste decomposition, nutrient cycling, pest control, flood control, climatic control (e.g., climate moderation, carbon sequestration), pollination of crops and native vegetation, and provision of recreational opportunities. Although not as well understood or appreciated as marketed nature-based activities, these nonmarketed ecosystem services, by all accounts, also are extremely valuable. Despite its value, biodiversity in LAC is increasingly threatened. The main underlying causes are high rates of population and economic growth that have spurred environmentally harmful human activities. Expansion of agriculture, urban areas, and coastal development have displaced natural biodiversity habitat. Agriculture, industry, and mining have generated pollution that threatens biodiversity directly and indirectly by degrading habitat. Agriculture, international trade, and human travel have introduced invasive species that have destabilized ecosystems. Unsustainable fishing, logging, and other extractive activities have depleted natural resources. And deforestation, agriculture, and industry have generated greenhouse gases that are changing the climate. LAC countries have used a wide variety of policies to stem these threats. However, a number of factors constrain their effectiveness. At least two have to do with inherent features



of biodiversity. First, as noted above, many of the services provided by biodiversity are left out of the marketplace, not by a conscious effort, but rather because they are not easily priced or traded. That dampens economic incentives to preserve biodiversity. For example, landowners are paid to produce crops and therefore have economic incentives to preserve the resources that enable them to do that. But they normally are not able to sell the ecosystem services their properties produce and thus do not have economic incentives to ensure continued provision of these ecosystem services. A related problem is that some of the value of biodiversity does not arise from human use. Examples are the value of preventing extinction of a species, apart from any value it would have to ecosystems or humans. Such nonuse values are difficult to measure or incorporate into management decisions. In addition to those somewhat subtle obstacles to biodiversity preservation are more obvious ones having to do with socioeconomic and institutional factors in many LAC countries. These include poverty and social problems that (rightfully) compete with biodiversity for scarce public funds, a lack of political will for strict enforcement of environmental regulations, weak regulatory institutions, and limited technical capacity. Given the enormous value of LAC biodiversity, increasing threats to that biodiversity, and the challenges to effective conservation, it is important to think carefully about how to prioritize policies for biodiversity conservation. Ecosystem management is one broad approach to biodiversity conservation policy. It emphasizes the inherent links among ecosystems, ecosystem services, and human activities: Although ecosystems provide valuable services to people, human activities can degrade their ability to supply these services. Spatial connectivity, location-specific factors and interactions further complicate these links. Ecosystem management is an integrated and adaptive approach to managing, using, and preserving ecosystems – including land, water, and living resources – so as to support both ecosystem services and human well-being. Despite considerable progress, significant barriers complicate putting the ecosystem management approach into practice, ranging from basic institutional capacity to lack of information on the ecological and economic processes underlying ecosystem services and their value to people.

1.2 Objectives This book has four specific objectives: 1 2 3 4

to describe the status of biodiversity in LAC, the main threats to this biodiversity, and the drivers of these threats; to identify the main policies being used to conserve biodiversity and assess their effectiveness and potential for further implementation; based on results from the effort to address the first two objectives, to propose five “lines of action”; and to identify the major actors engaged in biodiversity conservation in LAC.

1.3 Methods In compiling this book, we have relied mainly on secondary sources. Specifically, we drew on published and unpublished reports, journal articles, books, and book chapters, written in both English and Spanish. We augmented these secondary sources with two types of primary data. First, we used selected digital and printed primary data from the International Union of



Concerned Scientists (IUCN), the Economic Commission on Latin America (ECLAC), and the United Nations Food and Agriculture Organization (FAO), all described in the body of the book. Second, we conducted structured interviews of approximately one hour apiece with ten experts in LAC biodiversity representing multilateral and bilateral international cooperation institutions (the World Bank, United Nations Environment Program, US Agency for International Development, Centro Agronómico Tropical de Investigación y Enseñanza), nongovernmental organizations (The Nature Conservancy, Conservation International, and Forest Trends), academia (Georgia State University, Universidad Iboamericana), and a foundation (the Ford Foundation). Appendix 1 lists our interviewees. LAC is an exceptionally diverse region, not only in terms of the geography and biodiversity emphasized above, but also in terms of economics, politics, and culture. Hence, our discussion is necessarily somewhat broad-brush and general.

1.4 Organization The remainder of the book is organized as follows. Chapter 2 discusses the status and trends in LAC biodiversity. It is split into two subsections: one on terrestrial and freshwater systems and a second on marine and coastal ecosystems. Each of these subsections discusses the biophysical environment, status and trends in biodiversity by ecosystem type (mangrove, salt marsh, etc.), and finally threats to biodiversity by threat types (habitat loss, pollution, overexploitation, climate change, and invasive species). Chapter 3 discusses 17 policies (listed in the table of contents) used to conserve biodiversity in LAC. They are grouped into three categories: regulation and comanagement, marketbased policies, and a catch-all “other” category. Each subsection describes the policy, discusses status and trends in implementation in LAC, and highlights prominent implementation issues. For ten of 17 cases, we also review evidence on the policies’ effectiveness in a separate subsubsection. We omit such a discussion for the seven remaining policies for various reasons. A contributing factor in all seven cases is that, as discussed in Section 3.16, credible evidence on the effectiveness of biodiversity conservation polices is generally quite scarce. In addition, in some cases evidence of effectiveness is virtually nonexistent because the policy is either new in the region (mitigation offsets and banking) or has a predictable effect (wastewater treatment). Finally, in several cases the policy is cast so broadly – a necessary strategy given the scope of this effort – that evaluations of the policies as we have defined them do not exist (governance; targeting, data, and evaluation; fisheries management). In these instances, we have incorporated into the discussion citations to evaluations of more narrowly defined policies associated with the broad policy. For example, in the fisheries management subsection, we cite evaluations of catch share and transferable quota policies. As for the remainder of the book, Chapter 4 presents our five recommended lines of action. Finally, Chapter 5 briefly discusses our map of major LAC biodiversity actors, which is included in Appendix 3. For readers interested in specific LAC countries, Appendix 2 includes country-level tables, which complement subregion-level tables and statistics in the body of the book.


Status and trends

2.1 Terrestrial and freshwater systems 2.1.1 Biophysical environment Biophysically, the Latin America and Caribbean (LAC) region is exceptionally diverse. It supports 11 of Earth’s 14 terrestrial biomes, including wet and dry habitats; tropical, temperate, and desert systems; and forests, shrublands, and grasslands (Plate 2.1-1). It hosts five forest biomes (Olson et al. 2001; UNEP 2010c): •

Tropical and subtropical moist broadleaf forest is both LAC’s largest biome and its largest forest biome. It covers 44 percent of the region, including much of northern South America and parts of Central America and the Caribbean. Dominated by semievergreen and evergreen deciduous trees, this biome contains the highest levels of biodiversity of any LAC biome. Tropical and subtropical dry broadleaf forest is the next largest forest biome. It covers 6 percent of LAC, including parts of central South America, eastern Brazil, western Central America, and Cuba. It includes the world’s most endangered tropical and subtropical forests. Tropical and subtropical coniferous forest covers 2 percent of LAC. Found in Mexico, Guatemala, and the Caribbean, it also is critically endangered worldwide; 85 percent of the biome worldwide is in LAC. Temperate broadleaf and mixed forest also covers 2 percent of LAC. Found in the southern tip of South America and central Chile, it includes the Valdivian temperate forest, the second-largest temperate rainforest in the world, of which only 40 percent of the original cover remains. Mediterranean forests, woodlands, and shrubs are found in a small part of Baja California, Mexico, and in central Chile and have particularly high endemism.

LAC also contains several nonforest biomes (Olson et al. 2001; UNEP 2010c). •

Tropical and subtropical grasslands, savannahs, and shrublands are the second-largest biome in LAC and the largest nonforest biome. It covers vast areas of southern Brazil, Paraguay, Argentina, and all of Uruguay, as well as relatively small parts of Colombia and Venezuela. Temperate grasslands, savannahs, and shrublands, which differ from the biome just noted in temperature, rainfall, and tree cover, are mostly in Argentina.

Status and trends •

• •


Flooded grasslands and savannahs, which include wetlands, are scattered throughout LAC. One of the largest wetlands in the world, the Pantanal, is in central South America. This biome hosts many plant and animal species and is particularly important for migratory birds. Montane grasslands and shrublands are primarily confined to the Andean region, extending along the Pacific coast of South America. Deserts and xeric shrublands are scattered throughout LAC, including a large stretch of desert along central western South America, just east of the Andes in Chile and Peru. Another, exceptionally species-rich portion, the Caatinga, is on the eastern tip of Brazil and is the largest dry forest in South America. Mangroves, the final biome found in LAC, are described in the section on marine and coastal diversity.

2.1.2 Status and trends Species richness and overall threat levels LAC terrestrial biomes are remarkably biodiverse and host many species that are endemic. Although it constitutes only about 15 percent of Earth’s total land area, LAC supports 50 percent of its amphibians (i.e., half of known amphibians can be found in LAC, among other places), 41 percent of its birds, 35 percent of its reptiles, 33 percent of its mammals, and 32 percent of its vascular plants (UNEP 2010a; ICSU-LAC 2010). LAC includes six of the world’s 17 “megadiverse” countries, defined as countries hosting the largest numbers of endemic species: Brazil, Colombia, Ecuador, Mexico, Peru, and Venezuela (Sarukhán and Dirzo 2001; UNEP 2010b). Although no Caribbean countries are included in this list, the endemism of plants there is remarkable: half of Caribbean plant life is unique to the region. While LAC hosts significant species diversity, it also hosts some of the greatest numbers of threatened and endangered species: •

LAC contains five of the 20 countries with the highest numbers of endangered and threatened fauna and seven of the 20 countries with the most threatened plant species (UNEP 2010a). In LAC, 6,659 species – including 2,834 animal and 3,819 plant species – are among the 20,293 species worldwide that are included on the IUCN Red List of Threatened Species and are classified as extinct, extinct in the wild, critically endangered, endangered, or vulnerable. These include 1,131 amphibians, 458 birds, 366 mammals, and 237 reptiles. The Red List likely underestimates the species at risk, as it considers only those species for which status has been assessed. Nonetheless, 79 terrestrial species and 28 freshwater species in LAC are already extinct (IUCN 2011). As is generally the case worldwide, information on freshwater species richness and status in LAC is limited. This ecosystem is relatively poorly understood (ICSU-LAC 2010). As a result of species loss, the region’s genetic resources are being rapidly degraded. Approximately 40 percent of medicinal plant species in South America are threatened, and LAC has lost about 75 percent of its agricultural crop genetic diversity over the past 100 years (CBD 2010; UNEP 2010a).

Although the geographical distribution of both species richness and threats in LAC is varied, across the region most imminently threatened species are concentrated in nonremote fragmented habitats, principally in mosaic lands comprising both agricultural and natural


Status and trends

areas (Ricketts et al. 2005; Chomitz 2007). These are degraded habitats where species are under significant threat but have not yet gone extinct. Terrestrial ecosystems Despite LAC’s diversity of biomes, much of the discussion of its terrestrial biodiversity focuses on forests. Excluding Mexico, LAC forests cover almost 900 million ha, or 22 percent of global forests (Table 2.1-1; see Table A2.1-1 for country-level data).1 This forest covers half of LAC’s land mass. Seventy-six percent is primary forest, which is particularly rich in biodiversity – a much higher percentage than global forests. Four of the top ten countries in the world in terms of primary forest are in LAC: Brazil (35 percent), Peru (4 percent), Bolivia (3 percent), and Mexico (3 percent) (FAO 2011). Roughly half of LAC forests (460 million ha) are frontier forests – defined as tracts of continuous intact natural forest large enough to support viable populations of all biodiversity associated with the forest. LAC frontier forests account for 34 percent of the world’s total (ICSU-LAC 2010). The countries with the largest tracts of frontier forest, as a percentage of world total, are Brazil (17 percent), Peru (4 percent), Venezuela (3 percent), Colombia (3 percent), Bolivia (2 percent), and Chile and Argentina (2 percent). Chile and Argentina contain the world’s largest remaining tract of temperate frontier forest (ICSU-LAC 2010). A significant fraction of important forest ecosystems has already been lost because of conversion. The annual deforestation rate for LAC as a whole was 0.5 percent from 1990 to 2000 – more than double the global rate – and slightly lower, 0.4 percent, from 2000 to 2010 (Table 2.1-2; see Table A2.1-2 for country-level data). At the subregional level, for South America, the annual rate of deforestation was about half of 1 percent from 1990 to 2000 and from 2000 to 2010. Three of the 10 countries in the world with the largest annual net loss of forest area from 2000 to 2010 were in South America: Brazil, which lost 2.6 million ha, more than three times any other country; Bolivia, which lost 0.3 million hectares; and Venezuela, which also lost 0.3 million hectares (FAO 2011). It is important to note, however, that net loss of forest in South America actually declined between 2000 and 2010 – from a high of 4.4 million ha per year at the beginning of the decade to 3.6 million ha per year in 2010 – primarily because of reductions in deforestation rates in Brazil, which accounts for 60 percent of South America’s forests (FAO 2011).

Table 2.1-1 Forest area in LAC, by type and use, 2010 Region

Forest area (1,000 ha)

Primary forest area (1,000 ha)

Forest area (percentage of land area)

Primary forest area (percentage of forest area)

Production forest (percentage of forest area)

Planted forest (percentage of forest area)

Caribbean 6,933 Central America 19,499 South America 864,351 LAC* 890,783 World 4,033,060

205 4,482 624,077 628,764 1,102,382

30 38 50 49 31

4 23 76 71 36

28 19 14 14 30

11 3 2 2 7

Note *without Mexico. Source: FAO 2011.

Status and trends


Table 2.1-2 Trends in forest area in LAC, 1990–2010 Region

Annual change in total forest cover

Annual change in primary forest cover



1,000 ha/yr


Percentage 1,000 ha/yr

Caribbean 53 0.87 Central America –374 –1.56 South America –4,213 –0.45 LAC – –0.5 World –8,327 –0.20

50 –248 –3,997 – –5,211


Percentage 1,000 ha/yr

Percentage 1,000 ha/yr

0.75 –1.19 –0.45 –0.4 –0.13

–0.07 –0.98 –0.46 – –0.40

n.s. –54 –3,096 – –4,666

n.s. –74 –2,961 – –4,188

Percentage –0.02 –1.52 –0.46 – –0.37

Sources: FAO 2011; ECLAC 2011.

Although the total forest area lost was much smaller, the annual deforestation rate in Central America was the highest in LAC – 1.6 percent from 1990 to 2000 and 1.2 percent from 2000 to 2010. Forest cover in the Caribbean actually expanded in the past two decades,

Table 2.1-3 Trends in landcover by biome/region in LAC, 2001–2010 Biome/region

Woody vegetation1 2001 (1,000 ha)

Moist forests 607,521 Dry forests 123,332 Conifer forests 24,947 Temperate forests 14,685 Savannas/ 50,213 shrublands Pampas 4,648 Pantanal 5,458 Montane grasslands/ 1,737 shrublands Mediterranean 540 forests Deserts/xeric 66,661 shrublands South America 778,641 Mexico 95,119 Central America 20,142 Caribbean 5,841 LAC total 899,743

2010 (1,000 ha) 590,037 112,526 26,749 13,748 48,835

Mixed woody vegetation/ Agriculture/herbaceous plantations2 vegetation %3

−2.9 −8.8 7.2 −6.4 −2.7

3,476 −25.2 4,386 −19.6 1,633 −6.0

2001 (1,000 ha)

113,460 105,786 −6.8 28,510 32,142 12.7 15,168 15,197 0.2 7,397 9,205 24.4 74,680 78,676 5.4 15,752 2,712 3,923







751,391 104,728 19,604 6,080 881,802

−3.5 10.1 −2.7 4.1 −2.0

11,525 −26.8 2,992 10.3 3,201 −18.4 3,970

2001 (1,000 ha)

2010 %3 (1,000 ha)

208,608 233,446 11.9 56,019 63,937 14.1 9,891 8,469 −14.4 13,977 13,117 −6.2 152,288 151,657 −0.4 141,002 143,979 2.1 7,646 8,416 10.1 27,721 28,340 2.2



51,141 −18.7


257,796 249,571 −3.2 48,334 42,023 −13.1 15,831 15,176 −4.1 6,408 7,064 10.2 328,369 313,833 −4.4

Notes 1 closed-canopy (>80%) woody vegetation. 2 open-canopy (20-80%) woody vegetation and plantations. 3 decadal change. Source: Clark et al. 2012.

2010 %3 (1,000 ha)

1,133 −8.1 36,429


591,661 628,278 6.2 38,060 37,111 −2.5 13,900 15,048 8.3 9,329 8,486 −9.0 652,949 688,923 5.5


Status and trends

growing at more than three-quarters of 1 percent between 1990 and 2000 and between 2000 and 2010 (Table 2.1-2; see Table A2.1-2 for country-level data).2 Recent remote sensing data provide more detailed information on land cover change in LAC during the first ten years of the millennium, including how changes break down across different land covers, biomes, and subregions (Clark et al. 2012; Aide et al. 2013) (Table 2.1-3). As for types of land cover, from 2001 to 2010 LAC as a whole experienced a loss of 2.0 percent of its closed-canopy woody vegetation, a loss of 4.4 percent of its mixed woody vegetation and plantations, and an increase of 5.5 percent of its agriculture and herbaceous vegetation. In other words, to a large extent, forests and fragmented forests transitioned to agriculture. However, these trends vary greatly across subregions. For example, Central and South America experienced loss of woody vegetation and increases in agriculture/herbaceous vegetation, whereas the Caribbean and Mexico experienced the opposite. Land cover change also differed across biomes. Eighty percent of deforestation in the LAC from 2001 to 2010 occurred in three biomes – moist forest, dry forest, and savannas/shrublands – which not surprisingly also had the largest increase in agriculture/herbaceous vegetation (Aide et al. 2013). Among these three biomes, moist forests experienced the greatest area loss of woody vegetation. The Amazon Basin was the epicenter of this type of deforestation (Clark et al. 2012). Deforestation in dry forests largely occurred in Argentina, Paraguay, and Bolivia. Most of the deforestation in the savannah/shrubland biome occurred in South America (Aide et al. 2012). In contrast to these trends and reflecting regional heterogeneity, dry forests are increased in Mexico, Central America, and northern South America. There also were significant gains in woody vegetation in deserts and shrublands of Mexico, which may be associated with local climate changes (Clark et al. 2012). Just as the rate of forest loss varies across the region, so does the level of threat facing the wider diversity of terrestrial ecosystems (Plate 2.1-2). Many nonforested systems are highly threatened, and the threat level of the vast majority of the region is critical, endangered, or vulnerable. Global biodiversity hotspots are ecological regions that host extremely high biodiversity and are under significant threat because less than 70 percent of their original habitat remains (Mittermeier et al. 2005; Brooks et al. 2002). They help to identify high-priority conservation areas. LAC hosts nine of the world’s 34 biodiversity hotspots, including the California Floristic Province, Caribbean Islands, Madrean Pine-Oak Woodlands, Mesoamerica, Atlantic Forest, Cerrado, Chilean Winter Rainfall-Valdivian Forests, Tumbes-Chocó-Magdalena, and Tropical Andes (Plate 2.1-3). These regions include much of Central America and Mexico, the Caribbean, and the Pacific coast and the central Atlantic region of South America. Freshwater ecosystems Rainfall tends to be high or very high through much of LAC. Partly as a result, it has an abundance of freshwater resources – more than 30 percent of Earth’s available fresh water and roughly 40 percent of its renewable water resources (UNEP 2010a). LAC includes some of the world’s largest rivers (Amazon, Parana-Paraguay, Orinoco, Uruguay, MagdalenaCauca, and Usumacinta), wetlands (the Pantanal, Amazon wetlands, and southern South American temperate peatlands), lakes (Titicaca, Nicaragua, Managua, Maracaibo, and Chapala), and aquifers (the Guarani, Chaco, Puelche, and Valley of Mexico aquifers) (UNEP 2010c; ICSU-LAC 2010). The region contains a wide array of freshwater habitats, from large rivers and their deltas to rivers and streams in wet, xeric, and montane regions, flooded grasslands and savannahs, cold streams, bogs, swamps, and large lakes (Olson et al. 1998; Abell et al. 2008).

Status and trends


LAC contains 11 freshwater ecoregions considered globally outstanding in terms of biodiversity (Olson and Dinerstein 2002). Of these, however, seven have been classified as critical, endangered, or vulnerable based on the threats that they face: the Colorado River, Upper Parana Rivers and Streams, Brazilian Shield Amazonian Rivers and Streams, Greater Antillean Freshwater, High Andean Lakes, Mexican Highland Lakes, and Chihuahuan Freshwater systems (Olson and Dinerstein 2002). 2.1.3 Threats As discussed above, terrestrial and freshwater biodiversity in LAC is highly threatened in many regions. In this subsection, we first review IUCN data to shed light on which threats are most prominent in the region. We then discuss major threats one by one, focusing on identifying the activities and phenomena that spur them. Unfortunately, the terminology used in the relevant literature is confused – no common usage has yet emerged – so the same terms (threat, driver, factor, proximate cause, secondary cause, etc.) are used refer to different things (e.g., Angelsen 2007; UNEP 2007; Geist and Lambin 2002). Here we have adopted a simple lexicon. Threats are broad categories of phenomena that directly harm biodiversity. We use the most common typology, comprising five categories (e.g., Wilcove et al. 2000; Pereira et al. 2012): 1 2 3 4 5

habitat loss and degradation; overexploitation; climate change; pollution; and invasive species.

Drivers are activities, such as agricultural extensification and logging, that spur threats. Factors, in turn, are activities and phenomena, such as population growth, that spur drivers. Complicating this lexicon is the fact that the relationships among threats, drivers, and factors are inherently complex. Threats (e.g., climate change) can spur other threats (e.g., habitat loss), and drivers (e.g., logging) can spur other drivers (e.g., agriculture). Identifying and parsing the causal relationships that affect biodiversity is a research topic in itself and is beyond the scope of this book. Our aim is simply to identify the main activities and phenomena affecting biodiversity in LAC. Throughout, we discuss freshwater and terrestrial systems together. It is worth highlighting that although the relative importance of threats varies across regions, species, and ecosystems, habitat loss and degradation are believed to be the single most important threat to terrestrial biodiversity globally (Sala et al. 2000; Dirzo and Raven 2003; Pereira et al. 2012). Looking to the future, climate change and invasive species are going to be increasingly important (Sala et al. 2000; Pereira et al. 2012). For freshwater systems, invasive species may be more important than in terrestrial systems, and sedimentation from land-use change is a major threat (Sala et al. 2000). Within LAC, these same patterns generally hold (UNEP 2010a). IUCN Red List data The IUCN Red List, a global database of threatened and endangered species, catalogs the threats facing listed species (IUCN 2011). However, it does not use the common typology of five threats described above. Instead, it uses 11 “threat categories” that are combinations of


Status and trends

what we refer to as threats and drivers (Table 2.1-4; Salafsky et al. 2008). For each listed species, the IUCN identifies at least one of these categories as a “primary” threat. For many species, more than one threat is identified as primary. Table 2.1-4 shows the percentage of the species in LAC that the IUCN classifies as “vulnerable” to “extinct” for which each threat category is primary, by ecosystem type.3 For example, for terrestrial species in LAC classified as vulnerable through extinct, 19 percent face a primary threat from “residential and commercial development.” In addition, IUCN data include subcategories for each of the 11 categories in Table 2.1-4. For example, “livestock ranching” and “wood plantations” are subcategories under “agriculture and aquaculture.” Because the IUCN categories mix what we are calling threats and drivers, they do not provide a clear representation of causation. For example, numerous categories, including “residential and commercial development” and “agriculture and aquaculture,” can spur another threat category, “natural systems modifications.” Nevertheless, the IUCN – which to our knowledge has the only available data on threats by ecosystem and geographical region – provides a general sense of the importance of various activities and phenomena. For the terrestrial species on the list, the three most important (frequently listed) categories are “biological resource use” (particularly logging and wood harvesting), “agriculture and aquaculture” (especially annual and perennial nontimber crops and livestock ranching), and “residential and commercial development” (especially housing and urban areas). For the freshwater species on the list, the three most important are “agriculture and aquaculture” (particularly annual and perennial nontimber crops), “biological resource use” (particularly logging and wood harvesting), and “invasive and other problem species and genes.” We discuss the IUCN data on marine species (last column of Table 2.1-4) in the next section. Habitat loss and degradation This section discusses habitat loss and degradation first in the context of terrestrial ecosystems and then in the context of freshwater systems. Table 2.1-4 Percentage of IUCN Red List species (vulnerable – extinct) in LAC identifying each threat category as “primary,” by ecosystem type Threat category

Residential and commercial development Agriculture and aquaculture Energy production and mining Transportation and service corridors Biological resource use Human intrusions and disturbance Natural system modifications Invasive and other problem species and genes Pollution Geological events Climate change and severe weather Source: IUCN 2011.

Percentage of species Terrestrial



19 31 4 3 33 4 10 13 6 1 6

34 55 4 5 45 7 18 42 32 3 14

21 7 4 3 56 14 4 30 20 2 48

Status and trends



In terrestrial ecosystems, habitat loss and degradation threaten biodiversity in multiple ways. Habitat loss directly destroys organisms and habitat. It also fragments the habitat that is not destroyed, disrupting important ecological processes by limiting species dispersal and creating “edge” effects that cause diverse physical and biotic changes to the system (Saunders et al. 1991; Fischer and Lindenmayer 2007). Fragmentation can also hinder species’ abilities to adapt to anticipated climate change by preventing them from shifting their habitat range in response. Furthermore, habitat loss can disrupt ecosystem functions and inhibit the provisioning of valuable ecosystem services. For example, disruption of hydrological cycles and consequent reduced rainfall, a predicted result from massive loss of Amazon forest, can harm agriculture (Malhi et al. 2008). Similarly, the loss of natural habitat can lead to a decline in pollination services (Priess et al. 2007). Drivers of terrestrial habitat loss include agricultural extensification, logging, urban expansion, fire, and mining. As noted above, the relationships among these drivers are complex. Below, we discuss each in turn. Agricultural extensification. The main driver of habitat loss in LAC is clearing for agriculture. Between 1990 and 2008, cropland grew by 13 percent and pasture grew by 2 percent in LAC, with most of the growth in Latin America (Table 2.1-5; see Table A2.1-5 for country-level data). Pasture actually shrank by 6 percent in the Caribbean during this period. This trend also is evidenced by changes in agriculture and herbaceous vegetation in LAC from 2001 to 2010 (Table 2.1-3). This type of vegetation increased by 6.2 and 8.3 percent in South and Central America, respectively, but decreased by 2.5 and 9.0 percent in Mexico and the Caribbean. Across the LAC agriculture and herbaceous vegetation increased by 5.5 percent. The extensification of agriculture at the expense of native habitat has coincided with increases in agricultural production. Agricultural output (including forestry and fishing) grew 69 percent between 1990 and 2008 (Table 2.1-6; see Table A2.1-6 for country-level data). It grew much faster in Latin America (69 percent) than in the Caribbean (8 percent). The dynamics associated with agricultural expansion are diverse, complex, and regionally varied. Expansion of one form of agriculture (e.g., soybeans) may displace other forms (e.g., cattle ranching), which in turn relocate to the agricultural frontier and displace forests. Such complexity – along with limited land-use data – make it difficult to clearly identify and characterize the importance of the specific agricultural activities, such as export agriculture, ranching, and biofuels, that drive habitat loss (Chomitz 2007). Table 2.1-5 Area devoted to agricultural crops and pasture, by year and subregion (thousands of hectares) Region

Cropland 1990

Pasture 2000


Percentage 1990 growth, 1990–2008

LAC 149,975 161,778 169,747 13 Latin America 143,112 154,779 162,804 14 Caribbean 6,683 6,999 6,943 4 Source: ECLAC 2011.



Percentage growth, 1990–2008

533,941 548,543 542,318 2 529,064 544,103 537,747 2 4,877 4,440 4,571 –6


Status and trends

Table 2.1-6 Market value of agricultural output, by year and subregion (millions of 2005 US$)* Region




Percentage growth, 1990–2008

LAC Latin America Caribbean

91,011 89,572 1,438

115,728 114,041 1,687

152,830 151,273 1,556

68 69 8

Note *Includes fishing and forestry. Source: ECLAC 2011.

That said, significant evidence suggests that much of the observed agricultural expansion is due to large-scale, export-oriented commercial agriculture. In general, LAC agriculture has experienced a shift from basic food crops grown by small- and medium-scale producers and sold domestically, to agro-industrial export crops, such as soy, grown by large commercial operations (UNEP 2010b; IAASTD 2009). According to UNEP (2010a), almost half of deforestation in LAC results from expansion of commercial export agriculture. This trend has been especially prominent in northern Argentina, Bolivia, Paraguay, and Brazil (UNEP 2010c). In contrast, areas that recovered woody vegetation from 2001 to 2010 tended to occur in areas too steep or dry for modern agriculture (Aide et al. 2012). Roughly two-thirds of agricultural expansion in tropical Latin America between 1990 and 2000 replaced intact forests (Gibbs et al. 2010). There is strong geographic variation in this trend, with shrublands subject to greater conversion in less forested areas. Although soy expansion in southeastern Brazil replaced forest, pasture, and shrubland in roughly equal measures between 2000 and 2004, it was the main crop replacing intact forests (Gibbs et al. 2010; Morton et al. 2006). The growth in large-scale, export-oriented agriculture is reflected in production statistics. From 1990 to 2007, LAC soybean production grew 238 percent and sugar cane production 54 percent (Table 2.1-7; see Table A2.1-7 for country-level data). Livestock production is also a major driver of habitat conversion in some regions of LAC, particularly in parts of Venezuela, Brazil, Colombia, Ecuador, Guatemala, Nicaragua, Paraguay, Peru, and Bolivia (UN 2010). However, regional data paint a mixed picture. From 1990 to 2007, livestock production increased by 66.2 million head, reaching 392.3 million head, mainly in South and Central America (UN 2010). More recently, from 2005 to 2010, cattle production in South America has been fairly steady (FAO 2012), and the area devoted Table 2.1-7 LAC soybean and sugar cane production, by year and subregion (thousands of tons)* Region

Soybeans 1990

LAC 33,699 Latin America 33,699 Caribbean ––

Sugar cane 2000


Percentage 1990 growth, 1990–2007

57,339 57,339 ––

113,877 238 113,877 238 –– ––



Percentage growth, 1990–2007

492,419 536,026 760,719 54 382,067 471,686 720,799 88 110,352 64,340 39,920 –64

Note *Not all LAC countries included in regional totals. For details, see country-level data in Appendix 2. Source: ECLAC 2011.

Status and trends


to pasture has declined in the region as a whole and in both subregions since 2000 (Table 2.1-5). These statistics do not capture spatial trends in livestock production within countries, as when livestock production remains constant but is pushed to the agricultural frontier by competing land uses, and therefore may not be a good measure of its effect on habitat conversion. Biofuel production in Latin America has grown significantly over the past decades. Brazil, Colombia, and Paraguay are the leading producers of ethanol fuel in the region, and Colombia and Argentina have begun to export biodiesel derived from oil palm and soybeans (Janssen and Rutz 2011; Grau and Aide 2008; IAASTD 2009). The biofuel sectors in these countries are substantial. For example, 55 percent of sugar cane production in Brazil is used for ethanol and 35 percent of palm oil in Colombia is exported as biofuel. In 2009 Latin America produced 3.99 million terajoules of solid biofuels and 23 million tons of liquid biofuels (IEA 2012). Biofuels themselves may not be a leading driver of habitat loss and degradation, but cropland for feedstocks pushes cattle ranching and other forms of agriculture into the agricultural frontier. For example, soy production for biofuels has displaced a significant portion of ranchlands in Brazil (Janssen and Rutz 2011). Logging. Logging for timber and fuelwood is also an important source of deforestation and forest degradation (Chomitz 2007; Nepstad et al. 1999; TNC 2005). The main direct effect on habitat generally is degradation, since logging rarely entails clear cutting. For example, Asner et al. (2005) found in five Brazilian states that the area degraded by logging was greater than the area clear-cut. Selective logging damages remaining trees, other vegetation, and soils, causing erosion and altering hydrological processes, fire regimes, carbon storage, and biodiversity (Nepstad et al. 1999; Asner et al. 2005). Logging also can increase human access and can finance the clearing of land for agriculture. Therefore, it can be a factor spurring agricultural extensification (Chomitz 2007). Roads. Roads cause habitat loss, degradation, and fragmentation directly (Laurance et al. 2009). In addition, they are a major factor spurring clearing for agricultural and illegal logging: they lower the costs of these activities and improve access to unclaimed or loosely titled lands (Ibrahim et al. 2010; Chomitz 2007). Numerous studies have found that location near a road is an important predictor of deforestation (e.g., Kirby et al. 2006; Pfaff et al. 2007; for a review, see Chomitz 2007). Road building generally results from colonization, development projects, and natural resource extraction (TNC 2005; Chomitz 2007). Mineral and fossil fuel extraction. Mining and fossil fuel extraction contribute directly to habitat loss via deforestation and road building and by generating pollution that affects water, forests, and soils (see Section In addition to these direct effects, mining and other extractive activities also indirectly degrade habitat by increasing access to new areas, which in turn facilitates the expansion of agriculture and logging. In LAC, mining focuses on copper, coal, nickel, gold, silver, and sand. In the Caribbean, mining income grew from 10 to 20 percent of gross domestic product (GDP) from 1990 to 2010, and in the rest of LAC it has grown enough to maintain a constant 6 percent share of (growing) regional GDP (ECLAC 2011). Foreign investment in mining in LAC increased about 400 percent between 2000 and 2009, with an estimated $10 million invested in the sector annually (UNEP 2010b). The region has 10 percent of the world’s oil reserves and 14 percent of production (UNEP 2010b). Brazil, Mexico, and Bolivia are particularly dependent on the production and sale of fossil fuels. Fossil fuel production will probably increase, given that the industry is not well regulated and additional large fossil fuel reserves are likely to be discovered. Fire. In LAC fire causes forest loss and, more frequently, forest degradation (Nepstad et al. 1999). One cause of fires is accidental ignition, which tends to be particularly severe in


Status and trends

El Niño years (Cochrane 2003). But fire also is often used to clear land for agriculture and provide additional agricultural benefits, such as fertilization. Intentional fires, which can spread beyond their intended area, drive forest loss and degradation in remnant habitat patches and natural forest adjacent to agriculture, particularly in Central America (Chomitz 2007; TNC 2005). In LAC the extent and frequency of human-induced fire far exceeds historic rates, although rates vary by system and region (ICSU-LAC 2010). Between 2000 and 2004, about 3.3 million hectares of forest was lost to forest fires, especially in dry and semidry tropical forest ecosystems (UN 2010). Although fire is a natural part of some systems, including the grasslands and savannahs of Venezuela and Colombia, fire frequency in these systems has also increased because of humans (ICSU-LAC 2010). FRESHWATER ECOSYSTEMS

By changing water quantity, flows, and temperatures and reducing connectivity, dams and channelization inhibit the functioning of freshwater systems, thereby degrading freshwater biodiversity and the human activities that depend on it (Craig 2000; Bunn and Arthington 2002; Strayer and Dudgeon 2010; Pereira et al. 2012). For example, dams in the AraguaiaTocantins River basin of Brazil have blocked migration routes, reducing fish population in downstream areas by up to 70 percent (Craig 2000). Natural system modifications, including dams, water management, and water use, are cited as the primary threats to 18 percent of freshwater species ranked as vulnerable to extinct in the IUCN Red List (Table 2.1-4). Among natural freshwater ecosystems modifications, hydropower is probably the major driver of habitat loss and degradation. In South America, hydropower has grown rapidly in recent years as energy demands have increased. In the region as a whole, 60 percent of electricity produced is from hydro (Ray 2010). Water extraction and modification of surrounding terrestrial habitats also reduce freshwater biodiversity by causing wetland loss and affecting flows, water temperature, and water quality (Pereira et al. 2012). Because of the differences in water use and availability, the consequences of water extraction vary regionally. In LAC, surface water abstraction stress, defined as the ratio of water use (i.e., surface water withdrawn for domestic, crop, and livestock use) to water availability (measured as discharge by subbasin), is greatest in parts of Mexico, along the Pacific coast of South America, and in the Caribbean (Plate 2.1-4). FACTORS CONTRIBUTING TO HABITAT LOSS

Certain underlying factors strengthen the drivers of habitat loss discussed above. •

Population growth. Between 1950 and 2010, human population in LAC grew by more than 250 percent (ECLAC 2011). This rapid growth, along with comparable growth in other parts of the world, has increased demand for food, forest products, land, minerals, and energy, in turn spurring agriculture, logging, mining, oil drilling, and other activities that drive deforestation and habitat loss and degradation. Economic growth. GDP for LAC grew 87 percent between 1990 and 2010, and GDP per capita grew by 40 percent (ECLAC 2011). Again, this rapid growth, along with concurrent growth in other parts of the world, has increased demand for food, forest products, land, minerals, and energy, strengthening various drivers of habitat loss. It also has changed food consumption patterns. Demand for meat and milk products are increasing as the middle class grows (Delgado 2003; Ibrahim et al. 2010).

Status and trends •

• •


Globalization. Globalization has contributed to intermittently high prices for extractive export products, including crops and mineral resources, that in turn strengthen drivers of habitat loss, such as agricultural expansion and mining. According to a recent analysis, international trade is a driver of 30 percent of global species threats worldwide (not counting threats from invasive species) and is an even more important threat in developing nations (Lenzen et al. 2012). Technological change. Technological change can further the expansion of agriculture (Kaimowitz and Smith 2001). For example, advances in soils management led to expansion of soybean and eucalyptus cultivation in the Brazilian Cerrado and in the Mediterranean areas of central Chile. In addition, advances in ethanol production and use in Brazil have encouraged expansion of sugar cane production (ICSU-LAC 2010). Technology also can facilitate the clearing of forests by increasing clearing efficiency. Perverse incentives. Perverse economic incentives like agricultural price supports and other subsidies can promote deforestation, an issue discussed in Section 3.7. Biofuels policy. Latin American governments have begun to invest and plan for increasing biofuel production (Janssen and Rutz 2011). Biofuels are also discussed in Sections 3.7 and 3.15. Land distribution. Landless peasants and smallholders are important contributors to tropical deforestation and the expansion of the agricultural frontier (Aide et al. 2012). In some areas of LAC – particularly Central America – inequitable land distribution in combination with rising land prices has contributed to agricultural extensification by landless households, who often engage in small-scale shifting agriculture (TNC 2005; IAASTD 2009; Meyers and Tucker 1987; Kaimowitz 1996). Central America has the most inequitable land distribution in the world (Ferranti et al. 2004). For example, in Guatemala 65 percent of agricultural land is held by 3 percent of the population, and in Honduras 55 percent of land is held by 4 percent of the population. All told, more than one million rural families in Central America do not have access to land (TNC 2005). Also, a large proportion of land is held by absentee and latifundia owners (Meyers and Tucker 1987; IAASTD 2009). Exacerbating the situation, land prices in the region have been rising steeply near many urban areas, displacing poor producers living near cities (TNC 2005). Governance. Inadequate monitoring and enforcement of environmental regulations and implementation of land-use planning and other resource management policies contribute to deforestation, an issue discussed in Section 3.6. Invasive species Globally, invasive species – nonnative species that cause economic or ecological harm or damage human health – are a leading cause of rapid biodiversity loss (Kolar and Lodge 2001; Vitousek et al. 1996; Wilcove et al. 1998). They can reduce native biodiversity through competition for space or resources, predation, alteration of natural processes, or introduction of disease. Because many invasive plant species have weedy characteristics, they also can affect agriculture, forestry, fishing, and water supplies. Invasive species are listed as the primary threat to 13 percent of terrestrial species and 42 percent of freshwater species that are classified by IUCN as vulnerable to extinct (Table 2.1-4). Information on the presence of invasive species in this region is limited (ICSU-LAC 2010; Gardener et al. 2012). That said, what statistics do exist clearly show that invasives are a problem. Of the 613 highly invasive species registered in the Global Invasive Species Data


Status and trends

Base, 46 percent are registered as invasive in one or more countries or territories of LAC (ISSG 2012; ICSU-LAC 2010). These include plants, vertebrates, invertebrates, microorganisms, and fungi. Furthermore, 54 percent of the species in this database’s list of the “world’s 100 worst invaders” are present in LAC (ICSU-LAC 2010), and 42 percent of LAC’s freshwater species and 13 percent of terrestrial species in the IUCN Red List are at risk at least partly because of invasive species (Table 2.1-4). Examples of invasive animal species in LAC include European rabbit, which has spread to a large part of Chile, and decimates native vegetation; goats, which have destroyed large areas of native vegetation and encouraged establishment of exotic plant species on the Juan Fernández Islands; and the Canadian beaver, which in the Chilean Tierra del Fuego are causing serious flooding and conversion of huge amounts of forested land to sphagnum bogs. As for plant invasives, the introduction of buffelgrass (Pennisetum ciliare) into Guanica Dry State Forest in Puerto Rico provides fuel for frequent grass fires that suppress native fire-intolerant grasses, creating a feedback loop for further expansion of buffelgrass. Similarly, in the Bahamas, Old World climbing fern (Lygodium mycrophyllum) increases fire intensity by creating a fire ladder into tree canopies, thereby increasing the mortality of native species such as Caribbean pine (Pinus caribaea) (Burgeil and Muir 2010). In Central America, alien grasses prevent the regeneration of forests, increase fire risk, and can spread into forested areas (TNC 2005). Genetically modified organism (GMO) crops also are of concern. Although they may decrease the application of pesticides and herbicides, there is concern that they may spread into natural systems, hybridize with wild relatives, and potentially become invasive (TNC 2005; Hails 2000; Azadi and Ho 2010). GMOs are currently being used in Brazil, Canada, Colombia, Mexico, Honduras, Uruguay, and Paraguay (Traxler 2006). A number of drivers spur invasive species introduction. Invasive species are frequently introduced by humans, either intentionally as food, ornamentals, commercial production, or pets, or unintentionally, in ballast water, on imported products or packaging, and on vehicles and airplanes. Increasing international trade is likely to spur these activities (Gardener et al. 2012). There is also considerable concern that growing feedstocks for biofuel could lead to new invasions because the characteristics that make a species good for biofuel production may also make it a successful invader. Hence, trade, globalization, and biofuels policies are all factors contributing to invasive species threats. Following introduction, invasive species in LAC often go undetected and uncontrolled; as they spread, their economic and ecological damages increase. In addition, they can cross jurisdictional boundaries, necessitating management coordination for control (EpanchinNiell and Hastings 2010). Early detection and prevention efforts, which are thus far rare, become increasingly important as development and trade increase, but may be difficult to achieve in the region (Gardener et al. 2012). Climate change Climate change, including changing temperature, precipitation, and variability, is expected to have very significant adverse effects on LAC terrestrial and freshwater biodiversity and ecosystems (Strayer and Dudgeon 2010), leading to changes in species ranges, increases in disease and pest outbreaks, and unpredictable variability in populations and habitat conditions. These changes are likely to occur faster than some species can adapt, especially in conjunction with other stressors, such as habitat loss and fragmentation (Thomas et al. 2004; Pounds et al. 2006). The Intergovernmental Panel on Climate Change (IPCC) predicts that

Status and trends


climate change will cause the loss of 20 to 30 percent of LAC species that are at risk of extinction (UNEP 2010b). The vulnerability of biodiversity to climate change is likely to vary regionally because of differences in climate change and species’ ability to adapt. Existing models are too coarse to give reliable predictions of climate changes at the local level, but more warming is expected in the higher latitudes (ICSU-LAC 2010). In addition, over the past 50 years, mean annual precipitation east of the Andes in southern South America has increased by 35 percent, and water availability is predicted to decrease in Central America (ICSU-LAC 2010). In addition, the IPCC predicts that climate change will result in the complete melting of tropical glaciers in the Andes within 10–20 years (Bradley et al. 2006; UNEP 2010a). These glaciers make up 90 percent of the world’s total tropical glaciers and provide 10 percent of the world’s freshwater (Bradley et al. 2006; UNEP 2010a). Evidence also suggests that local climate change, mainly increases in precipitation, have resulted in the expansion of woody vegetation in the desert/xeric shrub biome of northern Mexico and northeast Brazil (Aide et al. 2012). As noted above, the effects of climate change tend to be exacerbated by other threats to biodiversity, such that species and ecosystems are susceptible to multiple interacting stressors. For example, habitat loss and fragmentation (e.g., from agriculture) are likely to hinder species’ ability to adapt through migration. In addition, reinforcing feedbacks between threats could exacerbate the effects of climate change. For example, climate change is expected to adversely affect both freshwater species and human use of freshwater, and the latter is likely to spur additional changes in water management that exacerbate the threat to freshwater biodiversity (Strayer and Dudgeon 2010). Anticipated interactions among deforestation, fire, and climate change in the Amazonian rainforest provide another example. In theory, interactions among these threats could lead to widespread forest dieback, which in turn could cause parts of the forest to enter a perpetual fire–drought cycle that ultimately converts forests to savannah-like ecosystems. The result would be net carbon loss, a decrease in precipitation, and a loss of biodiversity. According to some, the threat of such a scenario would increase dramatically if total deforestation in the Amazon exceeded 20 to 30 percent (currently, 17 percent has been lost) (CBD 2010; Malhi et al. 2008). Overexploitation Overexploitation of species occurs when the demand for the species is so high and/or the costs of harvest so low that there are private incentives to harvest the species faster than it reproduces. Various factors can contribute to the overexploitation of species, including high market prices and ill-defined property rights for the resource. The effects of overharvesting of terrestrial species are exacerbated by habitat fragmentation, which increases both species’ vulnerability and harvesters’ access. Overexploitation threatens many terrestrial and freshwater species in LAC and is an important threat to species worldwide (Dirzo and Raven 2003; Strayer and Dudgeon 2010; Pereira et al. 2012). Plant and animal harvests threaten around 7 percent of terrestrial species and 10 percent of freshwater species (ICSU-LAC 2010). Examples of terrestrial species that have been overexploited are pau-brasil (brazilwood), a woody member of the pea family, used for timber and to make bows for violins and other string instruments; Euterpe edulis (Arecaceae), a native palm tree of the Atlantic forest that is harvested for a food delicacy, the palm heart; the Chilean wine palm (Jubaea chilensis) of the monotypic endemic genus, exploited for a sweet honey product; and the Orinoco crocodile (Crocodylus intermedius) (ICSU-LAC 2010).


Status and trends Pollution Water pollution from domestic, industrial, and agricultural sources is a major threat to freshwater systems. Water pollution has a number of drivers. First, runoff from fertilized agricultural fields contributes heavily to nutrient loading that spurs dead zones affecting fish and other aquatic species (Pereira et al. 2012). Since the 1990s, fertilizer use has doubled in LAC (ECLAC 2011). Second, the use of pesticides and other agrochemicals, which has also grown, contributes to water pollution (ECLAC 2011). Third, soil erosion, particularly from deforested areas on steep slopes, contributes to sedimentation and degradation of freshwater systems, in addition to soil loss. Fourth, untreated sewage and industrial wastes, which we discuss in Section 3.5, exacerbate water pollution (Pereira et al. 2012). Finally, mining causes erosion and mercury contamination. Mercury contamination is a particular problem for small operations (UNEP 2010b) and has been shown to reduce biodiversity, including algal species richness (Giulietti et al. 2005). Factors contributing to pollution are varied. Because habitat loss and pollution share many of the same drivers – including agriculture, logging, and mining – many of the factors that contribute to habitat loss (discussed in Section also contribute to pollution. In addition, poor governance, specifically inadequate monitoring and enforcement of environmental regulations, is a contributing factor, an issue discussed in Section 3.6.

2.2 Coastal and marine systems LAC coastal and marine areas are rich in biodiversity (Chatwin 2007; Miloslavich et al. 2011) that has great economic importance, for example, via fishing and tourism (Bovarnick et al. 2010; FAO 2010; UNEP 2010), particularly in coastal areas, where the LAC population is concentrated (Cohen et al. 1997; Lemay 1998). Using data from the Center for International Earth Science Information Network, we estimate that more than 250 million people, about half (49 percent) of the entire LAC population, live within 100 kilometers of the coastlines (CIESIN 2007; see Table A2.2-S1 for country-level data). The concentration of populations on the coasts tends to be greatest in areas where coastal and marine areas are most productive and economically important. The close proximity of human populations threatens coastal and marine biodiversity through habitat loss and degradation, overexploitation, and introduction of invasive species (Halpern et al. 2008). Greenhouse gases and climate change pose a serious threat, especially in the long run (Behrenfield et al. 2006; Burkett et al. 2008). 2.2.1 Biophysical environment LAC contains seven main coastal and marine regions (Plate 2.2-1; Sullivan Sealy and Bustamante 1999): •

Tropical Northwestern Atlantic (Caribbean). The second-largest sea in the world, this region encompasses the Caribbean Sea, the Gulf of Mexico, and the Caribbean coasts of Central and South America. It has 22 independent states and 17 overseas territories, whose economies all depend heavily on marine environments. Frequently assigned top biodiversity conservation priority among all LAC marine regions, the broader region contains nearly 15 percent of the world’s coral reefs, including the Mesoamerican Reef, spanning 1,000 kilometers of coastline (Arrivillaga and Windevoxhel 2008). Although

Status and trends


the sea bottom in this region often drops to considerable depths relatively close to the coastline, coral reefs and seagrass beds populate the near-shore waters and are accompanied by mangroves on the coast, creating an ecologically and economically rich ecosystem. Three major South American rivers – Amazon, Orinoco, and Magdalena – carry enormous volumes of fresh water and sediments to the Caribbean Sea and have significant effects on the marine ecology. Southwestern Atlantic. The transition zone between the Tropical Northwestern and Southwestern Atlantic regions is dominated by mangroves and the vast deltas of the Orinoco and Amazon Rivers. Ocean currents flowing north from the mouth of the Amazon carry sediments and mixed fresh and sea water along the coasts of Suriname and Guyana. The Amazon River itself forms an estuary that stretches some 1,500 kilometers inland. South of the Amazon River’s zone, coral reef environments reappear along the northeastern and central coasts of Brazil. These reefs are relatively isolated geographically and thus host a number of endemic coral species. Warm-temperate Southwestern Atlantic. The coastal and marine areas in this region, including the southern coast of Brazil and coasts of Uruguay and Argentina, are highly productive marine systems that host significant marine biodiversity. Cold-temperate South America. The marine environment in this region is influenced by the cold currents flowing from the south. Around Cape Horn, one reaches the contoured coastline of southern Chile with its thousands of small islands, inlets, and fjords, where marine life, including cold-water corals, thrives. Southeastern Pacific. The cold and nutrient-rich currents in this region feed the highly productive fisheries of Chile and Peru. Cold currents flowing north meet the warmer waters off the coast of Ecuador, generating strong upwelling that gives rise to unusually high marine productivity along the Galapagos. Tropical Eastern Pacific. This region is dominated by mangrove forests and low-lying estuaries. In Central America, it is characterized by peninsulas, bays, and gulfs. Here, the wind patterns drive waters off the coast and draw in cold and nutrient-rich water from the deep sea, creating significant marine productivity. The entire Southeastern Pacific region is heavily influenced by El Niño, which changes ocean–atmospheric interactions and elevates sea water temperatures. Warm-temperate Northern Pacific. The Baja Peninsula, which forms one of the largest semienclosed marine basins in the world, dominates the coastal environments of this region. It hosts mangroves, coastal lagoons, coral reefs, and the rich biodiversity associated with them.

2.2.2 Status and trends Species richness As in most of the world, information on marine species richness in LAC is quite limited (ICSU-LAC 2010), a fact reinforced by the recently completed Census of Marine Life, the largest global research program on marine biodiversity, involving more than 2,700 scientists in some 80 countries (Costello et al. 2010; Miloslavich et al. 2011). Nonetheless, the census confirmed that marine environments in LAC display considerable biological diversity and a high degree of endemism (Miloslavich et al. 2011). Considerable variation in species richness exists within LAC. In South America, the Tropical East Pacific generally has a higher number of species than the Tropical West


Status and trends

Atlantic.4 Moreover, farther south, the Pacific side and the Humboldt current system are considerably richer than the Patagonian shelf at the same latitude (Miloslavich et al. 2011). The Caribbean region has a particularly high number of species – about 12,000 recorded marine species, more than for any other subregion in LAC (Miloslavich et al. 2010, 2011). The Caribbean region contains two global biodiversity hotspots, the insular Caribbean and the Mesoamerica hotspots, with particularly high richness of endemic species (Plate 2.2-2; Mittermeier et al. 2000; Myers et al. 2000). For example, the insular Caribbean hotspot has some 60 species of corals and some 1,500 species of fish, almost one-quarter of which are endemic (Heileman 2011). Concentration of species richness also varies by taxonomic group. The Tropical West Atlantic is a regional hotspot for fishes, whereas the Humboldt current along the southeastern Pacific coast is particularly rich for crustaceans. Brazil and Tropical Western Atlantic are rich in mollusks and Brazil has an especially rich environment in algae (Miloslavich et al. 2011). Coastal ecosystems Coastal ecosystems provide critical ecosystem services (Spalding et al. 2010; Barbier 1994; Barbier et al. 2008, 2011; Chmura et al. 2003). For example, they provide nursery habitats for fish, crustaceans, birds, and marine mammals (Twilley et al. 1996; Spalding et al. 2010; Mumby et al. 2004). Mangroves, salt marshes, and seagrasses in coastal areas store large amounts of carbon, often at relatively low opportunity cost (Donato et al. 2011; Siikamäki et al. 2012a, 2012b, 2013; Pendleton et al. 2012). In general, biodiversity in the coastal and marine ecosystems is highly dependent on the integrity of the coastal and near-coastal ecosystems (Miloslavich et al. 2011). Both globally and within LAC, coastal ecosystems are among the most threatened and rapidly disappearing natural environments (Valiela et al. 2001). As a result of continual conversion to other uses, coastal ecosystems have been degraded and their areas have been substantially reduced (e.g., FAO 2007; Spalding et al. 2010). Even the recent increase in coastal and marine protected areas has not stemmed the current tide of gradual degradation and disappearance of coastal ecosystems worldwide (Lotze et al. 2006; Halpern et al. 2008; Waycott et al. 2009; Spalding et al. 2010). Disturbances including outright loss are typically associated with human activities such as harvesting or the conversion of coastal ecosystems to agricultural, aquaculture, residential, and industrial uses. While these activities may have substantial economic benefits, they also typically diminish the capacity of ecosystems to support other beneficial services, such as carbon storage (Donato et al. 2011; Siikamäki 2012a, 2012b). Additionally, new pressures are emerging, such as sea-level rise, ocean acidification, and climate change more generally. MANGROVES

Mangroves are intertidal forests dominated by trees and shrubs that have evolved to thrive in anaerobic soils with varying levels of salinity (Spalding et al. 2010). LAC comprises more than 37,000 km2 of mangroves, one-quarter of the world’s total (Giri et al. 2010; Siikamäki et al. 2012a). Mangroves spread along the tropical coasts throughout the region (Plate 2.2-3; see Table A2.2-S2 for country-level data). South America’s 20,500 km2 of mangroves account for about 15 percent of the global total. Within South America, Brazil, Venezuela, Colombia, Suriname, and Ecuador possess particularly notable mangrove resources. The

Status and trends


rest of LAC contains roughly 15,300 km2 of mangroves, or about 11 percent of the global total. Major areas include the Yucatan Peninsula, Cuba’s northern barrier islands and western shores, and the southwestern stretch of the Bahamas. Roughly 40 percent of mangrove area in the Americas has been lost since 1980 (Valiela et al. 2001). Overall, the region loses more than 21,000 hectares, or 0.6 percent, of mangroves each year (Table A2.2-S2). The five countries with the greatest estimated total mangrove area lost annually are Mexico, Honduras, Venezuela, Colombia, and Panama, although several small island nations in the Caribbean have higher annual loss rates. SALT MARSHES AND SEA GRASSES

Like mangroves, salt marshes are often found in estuaries, deltas, and low-lying coasts that experience low wave energy. Salt marshes, however, have a greater latitudinal extent than mangroves (in the tropics, areas suitable to become salt marshes typically are subjugated by mangroves) and are dominated by herbaceous plants rather than trees. These plants are able to withstand high salinity and regular submersion due to high tides. Seagrass meadows comprise flowering species that provide shelter for aquatic animals and breeding grounds for various fishes (Kennedy and Bjork 2009). They also function as collection areas for sediments coming off the land and can provide important links between coral reefs and terrestrials systems like mangroves. Seagrass species occur in both tropical and temperate waters, often in relatively shallow waters because of their high light requirements for photosynthesis (Hemming and Duarte 2000). CORAL REEFS

The coral reef system along the Caribbean coasts of Mexico, Belize, Guatemala, and Honduras is the second-largest reef system worldwide. It provides habitat for an estimated 10,000 species and plays a critical role in the broader regional environment. Coral reefs are critically important to fisheries, and the productivity of fisheries is further enhanced by the coexistence of mangroves, seagrass beds, and coral reefs within the same ecosystem. Along with mangroves, corals reefs are among the most threatened and degraded ecosystems in LAC. Nearly two-thirds of coral reefs in this region are damaged by coastal development, sedimentation and other pollution, water acidification, and fishing activities. As a result, the Caribbean coral reefs have already lost almost one-third of their historical range, and it is estimated that another fifth will be lost in the absence of additional conservation measures in the next few decades (Gardner et al. 2003; TNC 2008; Sherman and Hemple 2009; USAID 2012). 2.2.3 Threats The leading threats to coastal and marine biodiversity in LAC are overexploitation of biological resources, especially overfishing; habitat loss and degradation; and pollution. Other significant threats are invasive species and climate change, especially ocean acidification and sea-level rise. In most cases, species and habitats are subject to two or more of these threats simultaneously (Miloslavich et al. 2011). Although the three main threats to coastal and marine systems are more or less the same across the LAC region, their relative importance and overall magnitude differ across and


Status and trends

within regions. For example, in Tropical East Pacific, the Census of Marine Life identifies the main threats to marine biodiversity as overfishing, climate change, habitat destruction, invasive species, and pollution. In the southern Southeastern Pacific, along the Humboldt current, overfishing and pollution are identified as the main threats. Within the Patagonian shelf, major threats to marine biodiversity are overfishing, habitat deterioration, urban and industrial pollution, and nonnative species (Miloslavich et al. 2011). The IUCN Red List, which as noted above is a global database of threatened and endangered species, catalogs the threats facing listed species (IUCN 2011). Table 2.1-4 shows the percentage of listed species in LAC in each threat category, by ecosystem type. Among marine species on the list, the most important (frequently listed) threats are “biological resource use,” “agriculture and aquaculture,” “invasive and other species,” and “residential and commercial development.” We next discuss the main threats and their drivers. Habitat loss and degradation As in terrestrial systems, habitat loss and degradation in coastal and marine systems threaten biodiversity through direct destruction, habitat fragmentation, disruption of ecological processes, and impairment of ability to adapt to climate change. It is a major problem for coastal systems. As noted above, coral reefs in the region have lost a considerable portion of their historical range, and the majority of the remaining coral reefs are degraded and threatened. The area of mangroves is rapidly shrinking, especially in the Caribbean and Central America (Siikamäki et al. 2012a, 2012b; Giri et al. 2010). Habitat losses often result from drivers of coastal development, such as tourism or agricultural development. Climate change also is a significant driver; for example, sea-level rise due to climate change can reduce the amount of coastal habitat. Ocean acidification due to increased concentrations of carbon dioxide is detrimental to marine biodiversity, especially shellfish and corals, which are also damaged by the rise of sea water temperature. Another source of habitat degradation involves destructive fishing practices, such as bottom trawling, that damage coral reefs and seagrass beds. Overfishing can also reduce the abundance of herbivores, which in turn can damage corals by the overgrowth of algae (Heileman 2011). Pollution As noted above, approximately half of LAC’s population lives in coastal areas, and concentration of populations in coastal areas is highest in places like the Caribbean, where coastal and marine environments are highly productive and economically important. The resulting pollution from agricultural, industrial, and urban sources is a significant driver of biodiversity loss. Agricultural nonpoint source pollution is often a leading cause of water pollution in LAC, as it is worldwide. Municipal wastewater is a particularly pressing problem in LAC, where some 85 percent of all municipal waste is discharged without any treatment (PAHO 2001). This type of pollution will worsen in the future because human population in LAC coastal areas is rapidly growing. Increased sediment flow due to land-use changes and industrial processes also contribute to water pollution. Drivers – particularly of agricultural intensification and expansion – are discussed in more detail in the sections above. Both agricultural pollution and municipal waste release large amounts of nitrates and phosphates into aquatic systems, where they can cause algal blooms that choke off oxygen. Massive blooms, known as red tides, seriously damage ecosystems and fisheries and are an

Status and trends


especially serious problem in the Gulf of Mexico surrounding the areas that receive water flows from the Mississippi River. However, red tides have also occurred in South America (WRI 2008). In addition, coastal and marine areas also suffer from other types of chemical pollution, including from persistent organic pollutants (POPs). POPs accumulate in ecosystems, creating long-term hazards to ecological and human health. For instance, polychlorinated biphenyls (PCBs) can harm human health through consumption of fish and seafood. Animals, especially those high up in the food chain, such as birds of prey, also suffer from the bioaccumulation of contaminants, especially through reduced reproductive ability (UNEP 2006). Overexploitation Overexploitation occurs when the harvesting of a particular species overwhelms its ability to regenerate, leading to a decline in population. At the extreme, exploitation can lead to a natural resource stock collapse. Rebuilding a collapsed stock can take a very long time and in some cases is altogether impossible. Overexploitation, especially overfishing, is a particularly serious problem in marine ecosystems and is a significant driver of biodiversity loss. Production (landings) in LAC capture fisheries, including both coastal and marine operations, has exhibited the same temporal pattern over the past several decades as in the rest of the world: despite increases in fishing capacity (i.e., number and capacity of fishing vessels) production has reached a plateau or has already passed it (FAO and World Fish Center 2008; Worm et al. 2009; FAO 2010; Salas et al. 2011). Total landings from coastal fisheries in LAC rapidly increased throughout the 1960s and then collapsed in the early 1970s (Figure 2.2-1; see Table A2.2-S3 for country-level data). A gradual increase of landings followed until around the mid-1990s, after which landings have fluctuated considerably with a slight downward trend. Latin American (versus LAC) fisheries exhibited the same overall trend. In the Caribbean, landings from the coastal fisheries peaked in the mid-1980s and have since steadily declined (Figure 2.2-2). Coastal fishing pressure in LAC is relatively high throughout the region (Plate 2.2-4). It is particularly high in the Caribbean, the coasts of Brazil, Ecuador, Peru, and Chile, and the Pacific coast of Central America (Stewart et al. 2010). Fishing pressure in marine areas is also heavy. Its geographical distribution is roughly the same as that of coastal fishing pressure (Halpern et al. 2008). These trends clearly indicate that fisheries in LAC are not on a sustainable resource use path. Several policy-relevant factors have contributed to this problem (Swan and Gréboval 2004; Salas et al. 2011): • • • • • • •

high and growing demand for limited fishing resources (see Section 3.4); inappropriate economic incentives, including subsidies that encourage overcapacity (see Section 3.7); lack of knowledge about fisheries and the associated uncertainties (see Section 4.5); lack of governance (see Section 3.6); poverty and lack of alternatives in coastal communities; interactions of the fishery sector with other economic sectors and their environmental effects; and stock fluctuations due to natural causes.


Status and trends 30

Landings (millions of tons)






0 2008 2006 2004 2002 2000 1998 1996 1994 1992 1990 1988 1986 1984 1982 1980 1978 1976 1974 1972 1970 1968 1966 1964 1962 1960 1958 1956 1954 1952 1950 Year

Figure 2.2-1 Landings from coastal fisheries in LAC. Source: compiled using data from ECLAC 2011. Climate change Global climate change and the atmospheric accumulation of human-generated greenhouse gases are a substantial threat to marine biodiversity in LAC. Although species have certain capacity to adapt to environmental change, climate change likely will occur more rapidly than most previous, natural climate shifts. Sea-level rise, ocean acidification, and alteration of ocean currents and precipitation regimes could potentially have serious consequences for coastal and marine biodiversity, including habitat loss (e.g., coastal inundation, coral reef loss), expansions or contractions of natural species ranges, increases in disease transmission and pest infestations, and unpredictable fluctuations in populations and habitat conditions. Within LAC, sea-level rise and ocean acidification are seen as the main climate change-related threats to coastal and marine biodiversity (Miloslavic et al. 2011). The adaptive power of some species will likely be overwhelmed by these new pressures, especially if combined with fragmentation, decreased connectivity of habitats, and other stresses that may create additional barriers to adjustment (Thomas et al. 2004; Pounds et al. 2006). For example, mangroves may be able to expand their range in response to sea-level rise but require favorable topography and absence of other barriers, such as agriculture, roads, and developed areas. If mangroves cannot migrate inland quickly enough, they will gradually be fragmented and lost (Spalding et al. 2010).

Status and trends




Landings (thousands of tons)







0 2008 2006 2004 2002 2000 1998 1996 1994 1992 1990 1988 1986 1984 1982 1980 1978 1976 1974 1972 1970 1968 1966 1964 1962 1960 1958 1956 1954 1952 1950 Year

Figure 2.2-2 Landings from coastal fisheries in the Caribbean. Source: compiled using data from ECLAC 2011. Invasive species Invasive species form an increasing threat to LAC marine biodiversity. The Caribbean is considered particularly problematic in this regard. Lopez and Krauss (2006) compile information on more than 100 marine invasive species. The Census of Marine Life (Miloslavich et al. 2010) finds altogether 44 introduced species. The red lion fish, which can seriously disrupt coral reef communities, is one of the best-known invasive species. Other nonnative fish, oysters, sharks, algae, mussels, and other species have also established themselves. Extensive marine transport in the region contributes to the introduction of nonnative species and therefore is a driver of biodiversity loss. Nonnative species invading the Caribbean mostly originate from the Mediterranean and the Indo-Pacific region (Heileman 2011).

Notes 1 FAO groups Mexico with the United States and Canada in North America. 2 ECLAC statistics, which are used only for the LAC row in Table 2.1-2, but for all of Tables 2.1-4, 2.1-5, and 2.1-6, exclude the following countries typically included in LAC: Anguilla, Aruba, British Virgin Islands, Cayman Islands, Curacao, Falkland Islands, French Guiana, Guadeloupe, Martinique, Montserrat, Navassa Island, Puerto Rico, St. Bethelemy, St. Martin, Turks and Caicos Islands, and Virgin Islands.


Status and trends

3 IUCN designates species as (ranging from most to least threatened) extinct, extinct in the wild, critically endangered, endangered, vulnerable, near threatened, or least concern. Species also may be categorized as data deficient or not evaluated if data are lacking or the species has not been assessed. 4 The naming and delineation of marine regions in this and the following section somewhat differ from the section above, but follow the original source (Miloslavich et al. 2011).




3.1 Terrestrial protected areas 3.1.1 Description Protected areas (PAs) are arguably the most common and highest-profile conservation policy worldwide. According to the IUCN, a PA is “a clearly defined geographical space, recognized, dedicated and managed through legal or other effective means to achieve the longterm conservation of nature with associated ecosystem services and cultural values” (Dudley 2008). PAs are used for both terrestrial and marine conservation. The IUCN defines six main categories of PAs (Dudley 2008): • • • • • • •

Category Ia: strict nature reserve: a PA managed mainly for science. Category Ib: wilderness area: a PA managed mainly for wilderness protection. Category II: national park: a PA managed mainly for ecosystem protection and recreation. Category III: natural monument: a PA managed mainly for conservation of specific natural features. Category IV: habitat or species management area: a PA managed mainly for conservation through management intervention. Category V: protected landscape or seascape: a PA managed mainly for landscape or seascape conservation and recreation. Category VI: managed resource: a PA managed mainly for the sustainable use of natural ecosystems.

These types are often grouped into two metacategories: “strictly” protected areas, comprising categories I–IV, and “mixed-use” areas, comprising categories V and VI (Nelson and Chomitz 2011; Jenkins and Joppa 2009). The former PAs entail prohibitions on productive activities, such as logging and agriculture, whereas the latter allow them subject to “sustainable” management. 3.1.2 Status and trends Protected area coverage standards For at least 20 years, the international conservation community has had a goal of protecting 10 percent of terrestrial and marine territory. This goal pertains to the world as a whole,



individual countries, and biomes within countries. A 10 percent goal for terrestrial territory was first proposed at the 1982 World Parks Congress in Bali, Indonesia, and was subsequently incorporated into numerous high-profile international declarations, including the 1992 Convention on Biodiversity, signed by 167 countries at the United Nations Conference on Environment and Development in Rio de Janeiro (Naughton-Treves et al. 2005). Coverage by region As a region, LAC has led the developing world in PA coverage for several decades. As early as 1990, at the start of a period of rapid global expansion, LAC PAs encompassed 10 percent of the region’s terrestrial area versus 9 percent for the entire developing world (Table 3.1-1; see Table A3.1-1 for country-level data). By 2010 they encompassed 20 percent of the region’s terrestrial area versus 13 percent for the entire developing world. Most of the expansion of PAs in LAC over the past 20 years occurred in Latin America, not the Caribbean. By 2010, 20 percent of terrestrial area in Latin America was protected versus just 11 percent in the Caribbean (Table 3.1-1). As of 2010, only five of the 21 countries in Latin America had protected less than 10 percent of their terrestrial area. By contrast, in the Caribbean, 12 of 26 countries had protected less than 10 percent of their terrestrial area (see Table A3.1-1). Coverage by country At the country level, several LAC nations stand out as leaders in establishing PAs: Venezuela (54 percent of terrestrial area), French Guiana (48 percent), Nicaragua (37 percent), Guatemala (31 percent), Belize (28 percent), and Brazil (26 percent) (Table A3.1-1). PA coverage grew particularly rapidly between 1990 and 2010 in Brazil, Mexico, and Peru. Notwithstanding this impressive regional growth, several Latin American countries still are far from meeting the 10 percent goal: Uruguay (0.3 percent), El Salvador (1 percent), Guyana (5 percent), and Paraguay (5 percent) (Table A3.1-1). Coverage by type of protection In addition to leading the developing world in coverage of PAs, LAC also leads in establishing multiuse PAs – those in IUCN categories V and VI – as well as those targeting areas inhabited by indigenous groups (Table 3.1-2). Table 3.1-1 Percentage of terrestrial and marine area protected, by region* Area

World (outside Antarctica) Developed regions Developing regions Latin America and Caribbean Latin America Caribbean *Marine













8.8 8.7 8.8 9.7 9.7 9.2

11.3 10.7 11.7 15.3 15.4 9.9

12.7 11.6 13.3 20.3 20.4 11.2

3.1 5.9 1.0 2.7 3.3 1.1

5.2 8.5 2.9 8.9 11.8 1.5

7.2 11.5 4.0 10.8 14.3 2.2

8.1 8.3 7.9 9.0 9.3 3.3

10.6 10.4 10.6 14.7 15.1 3.8

12.0 11.6 12.2 19.3 19.9 4.6

area consists of territorial waters (within 12 nautical miles of coast). Source: IUCN and UNEP-WCMC 2011.



Table 3.1-2 Percentage of terrestrial protected areas, by type and region, 2000 Type





Strict (I–IV) Multiuse (V–VI) Unknown Indigenous Other

26 33 8 31 1

37 6 53 0 4

54 22 23 0 1

31 29 15 23 2

Source: adapted from Nelson and Chomitz 2011. Coverage by biome Despite relatively high levels of terrestrial PA coverage in LAC, certain biomes are underrepresented at both regional and national levels. In LAC, as in most of the developing world, PAs have focused mainly on tropical forests and, to a lesser extent, mangroves and temperate broadleaf forests (Table 3.1-3). By contrast, less than 2 percent of (i) temperate grasslands, savannahs, and shrublands and (ii) Mediterranean forests, woodlands, and scrub have been protected. 3.1.3 Issues Coverage gaps As discussed above, although PA coverage in LAC is high relative to other developing regions, significant gaps persist both at the country and the biome levels. Such gaps have been widely cited as justification for expanding PA coverage in LAC (Flores 2010). Of particular concern is the underrepresentation in PAs of biomes other than tropical forests, Table 3.1-3 Percentage of terrestrial protected areas, by biome, region, and year Biome

All Tropical and subtropical moist broadleaf forests Tropical and subtropical dry broadleaf forests Tropical and subtropical coniferous forests Temperate broadleaf and mixed forests Temperate coniferous forests Boreal forests and taiga Tropical and subtropical grasslands, savannahs, and shrublands Temperate grasslands, savannahs, and shrublands Flooded grasslands and savannahs Montane grasslands and shrublands Tundra Mediterranean forests, woodlands, and scrub Deserts and xeric shrublands Mangroves Sources: adapted from Brooks et al. 2004; Jenkins and Joppa 2009.







12 18 9 6 9 25 10 13 5 18 15 15 6 10 16

13 21 8 7 11 25 9 13 4 20 25 17 7 9 21

16 24 9 8 25

20 32 9 8 29

8 3 10 13

11 2 15 14

1 8 26

1 9 37



specifically (i) temperate grasslands, savannahs, and shrublands and (ii) Mediterranean forests, woodlands, and scrub. Although these biomes are generally not as rich in biodiversity as tropical forests, they are often under greater threat and provide important ecological services such as hydrological services or carbon sequestration (Hoekstra et al. 2004; Kareiva and Marvier 2003). Global studies aimed at identifying geographic areas where additional coverage is most urgently needed tend to highlight LAC countries, among others (Brooks et al. 2004; Jenkins and Joppa 2009; Rodrigues et al. 2004). For example, Rodrigues et al. (2004), who take into account threat as well as biome coverage gaps, prioritize the Andes and neighboring lowland Pacific forest of the Chocó and Tumbes, the Atlantic forest, the Caribbean, and Central America. Size and fragmentation Closely related to concerns about PA coverage gaps in LAC are concerns that many PAs in the region are too small and fragmented to effectively stem the loss of biodiversity and ecosystem services. Some research suggests PAs must encompass at least 10,000 hectares to support far-ranging species and ecosystem processes like natural fire regimes (NaughtonTreves et al. 2005). In addition, some have argued that only relatively large PAs create buffers against encroachment that are large enough to effectively stem land-use change (DeFries et al. 2005). Also, corridors between PAs can facilitate climate adaptation (Magrin et al. 2007; Yadvinder et al. 2008). Such arguments have given rise to calls for consolidating, connecting, and expanding PAs. Management Like PAs in other developing regions, those in LAC are often poorly managed. Common problems are weak monitoring and enforcement of legal protections, gaps and inconsistencies in legal and regulatory underpinnings (e.g., definitions of boundaries and property rights), and poor coordination with local communities and other governmental agencies (Brandon et al. 1998; Dudley and Stolton 1999; Leverington et al. 2010). In extreme cases, which are not uncommon, PAs are “paper parks” that have little or no formal management (Bonham et al. 2008). Drivers of inadequate PA management include lack of political will for stringent environmental and natural resource management, conflicts with economic development, and inadequate fiscal and human resources. Financial resources As in the rest of the developing world, financial resources devoted to PAs in LAC are far short of what is needed for effective management (Bruner et al. 2004; Bovarnick et al. 2010b). On average, LAC governments allocate less than 0.01 percent of GDP to PAs and 1 percent of national environmental budgets, which amounts to $1.18 per PA hectare (see Table A3.1-S1 for country-level data). These allocations, along with funding from international sources, cover only 54 percent of the basic financial needs of existing PAs and 34 percent of what would be needed for “optimal” financial management (Bovarnick et al. 2010b). An additional $317 million per year is needed to cover basic needs, and $700 million per year is needed to ensure optimal management (Bovarnick et al. 2010b). Brazil and Mexico account for more than half of this funding shortfall (Table 3.1-4). Even more funding



Table 3.1-4 Protected area management costs and financial gaps in selected LAC countries, 2010 ($) Country

Financial needs (costs) Current funding Basic needs

Argentina Bolivia Brazil Chile Colombia Costa Rica Cuba Dominican Republic Ecuador El Salvador Guatemala Honduras Mexico Nicaragua Panama Paraguay Peru Uruguay Total

31,309,584 5,102,653 133,415,026 9,194,339 18,026,595 29,645,948 14,587,030 10,380,071 3,977,600 3,803,223 8,339,504 4,122,552 80,214,239 5,314,245 9,506,948 1,240,665 13,067,100 816,000 382,063,322

Financial gaps (costs – current funding) “Optimal”

Basic needs


39,512,820 60,366,666 8,203,236 29,057,082 5,374,940 9,000,000 272,287 3,897,347 302,573,314 471,731,602 169,158,288 338,316,576 17,974,193 26,754,046 8,779,854 17,559,707 25,150,153 42,755,260 7,123,558 24,728,665 31,934,374 44,000,000 2,288,426 14,354,052 21,639,821 36,787,695 7,052,791 22,200,665 22,574,294 27,974,294 12,194,223 17,594,223 6,730,054 14,040,147 2,752,454 10,062,547 4,445,738 7,557,755 642,515 3,754,532 16,118,443 27,401,353 7,778,939 19,061,849 6,618,629 11,251,670 2,496,077 7,129,118 120,321,358 160,428,478 40,107,119 80,214,239 19,546,456 43,321,382 14,232,211 38,007,137 19,880,360 33,796,612 10,373,412 24,289,664 9,700,000 19,500,000 8,459,335 18,259,335 25,172,664 41,842,414 12,105,564 28,775,314 3,409,002 4,355,947 2,593,002 3,539,947 698,676,613 1,082,865,321 316,613,291 700,801,999

Source: Flores 2010.

would be needed to expand the existing network of PAs. An additional $22 million per year would be required to expand the PAs network in Brazil, Bolivia, Chile, Colombia, Ecuador, Peru, and Venezuela to fill the type of coverage gaps described above (TNC 2007). In general, financial management of LAC PA systems is poor (Bovarnick et al. 2010b). An effort to evaluate management among 19 Latin American countries found that the average score was just 60 percent of the maximum possible score. The 19 countries were placed in three groups based on their scores (Bovarnick et al. 2010b): • • •

Strong financial planning (score higher than 50 percent): Costa Rica, Cuba, Colombia, Argentina. In need of strengthening (score of 30–50 percent): Mexico, Panama, Honduras, the Dominican Republic, Ecuador, Bolivia, Peru, Venezuela. In need of substantial strengthening (score below 30 percent): Belize, El Salvador, Guatemala, Nicaragua, Brazil, Paraguay, Chile, Uruguay.

On average, 60 percent of funding for PAs from these 19 countries comes from government monies, 15 percent from international cooperation, 14 percent from site-based sources such as entrance fees, and 11 percent from other sources (Flores 2010). Local communities The relationship between PAs and the people living in and around them has generated considerable concern. PAs can both impose costs on and provide benefits to local communities.



On one hand, rural households rely on local landscapes for hunting, collecting biomass for fuel and fodder, logging, agroforestry, and shifting agriculture, and they use local water bodies for fishing and other extractive activities. By their nature, PAs restrict such activities. Plentiful anecdotes suggest that such restrictions can impose costs. For example, the creation of a national park in Nicaragua ultimately led to the economic and social disintegration of the resident communities (Kaimowitz et al. 2003). Such striking case studies no doubt motivated the 2004 World Parks Congress to declare that “many costs of protected areas are borne locally – particularly by poor communities” (Ferraro 2008). On the other hand, PAs can generate significant local economic benefits by boosting ecotourism, attracting outside investment in roads and other infrastructure, and protecting natural resources that provide critical ecosystem services (Bovarnick et al. 2010a). For example, according to Mexico’s Ministry of the Environment, the country’s PAs are visited by six million people every year, generating $300 million in economic benefits, 90 percent of which accrues to local communities (Elvira-Quesada 2007). Empirical evidence of the net effect of PAs on local communities in LAC is mixed and suggests, perhaps not surprisingly, that site-specific characteristics are determinative (Canavire-Bacarreza and Hanauer 2013; Miranda et al. 2013; Ferraro et al. 2011; Ferraro and Hanauer 2011). 3.1.4 Effectiveness We probably know more about the effectiveness of PAs than about any other conservation policy. Dozens, if not hundreds, of studies have been published, mostly focusing on a single terrestrial PA or a small collection of them in a single country (Geldmann et al. 2013; Miteva et al. 2012; Joppa and Pfaff 2010b; Nagendra 2008; Naughton-Treves et al. 2005). A considerable literature uses remotely sensed (satellite and airplane), survey, and other quantitative data to gauge the effect of PAs on land-cover change. However, almost all of these studies have methodological problems that tend to bias their estimates of effectiveness upward (Joppa and Pfaff 2010b). Specifically, most do not correct for the tendency of terrestrial PAs to be sited in remote areas with minimal deforestation pressure (thereby conflating the effect on land-cover change of the PAs and their siting), and most do not allow for the possibility that PAs displace extractive activities to adjacent areas. A newly emerging literature uses statistical techniques to correct for these problems (Blackman 2013a). The results of the older, flawed studies and newer, more rigorous ones are mixed. The most common finding, including that of the handful of recent studies that correct for the methodological problems noted above, is that PAs stem land-cover change, although the size of this effect is often modest. For example, global studies comparing rates of tropical deforestation or fire incidence (a proxy for deforestation in tropical areas, where virtually all large-scale fires are deliberately set to clear forest) inside PAs and in nearby buffer areas find that the former rates are lower (DeFries et al. 2005; Wright et al. 2007), although it is important to point out that the finding may be at least partly driven by leakage. Evaluations based on interviews with local stakeholders, such as park managers, generate similar conclusions, as do meta-analyses of park- and regional-level studies (Bruner et al. 2001; Nagendra 2008; Naughton-Treves et al. 2005; Geldmann et al. 2013). Particularly compelling are recent global studies that control for the methodological issues described above (Joppa and Pfaff 2010a; Nelson and Chomitz 2011). For example, Nelson and Chomitz (2011) find that in LAC, strict protection reduces fire incidence by 3–4 percentage points compared with what it would have been absent protection; multiuse protection reduces it by 5–6 percentage points; and PAs in indigenous areas reduce it by 16–17 percentage points. These global



studies include parks with a variety of levels of funding and management and do not tell us much about how these factors correlate with effectiveness. Notwithstanding those positive evaluations, a significant number of studies find that PAs are decidedly ineffective in stemming land-cover change, even though many of them have methodological problems that tend to bias their estimates of effectiveness upward (Fuller et al. 2004; Liu et al. 2001; Viña et al. 2007; Curran et al. 2004; Gaveau et al. 2009). Some of the more rigorous studies that attempt to control for these problems echo these negative results (Blackman et al. 2010; Cropper et al. 2001). Hence, it is reasonable to conclude that although PAs can be effective in stemming landcover change, their effectiveness depends on site-specific factors such as land-cover change pressures and PA management and funding. Indeed, studies show that management and funding are directly correlated with PA effectiveness (Nolte et al. 2013; Bruner et al. 2001, 2004; Wilkie et al. 2001). An emerging topic in the literature is the relative effectiveness of strictly protected PAs versus mixed-use PAs. Recently, a number of rigorous studies that control for nonrandom siting of such PAs have appeared. Here, too, the evidence is mixed: several studies find that strict protection outperforms mixed-use protection (Nolte et al. 2013; Soares-Filho et al. 2010; Ferraro et al. 2013) while others find the opposite (Nelson and Chomitz 2011; Pfaff et al. in press; Blackman 2013b). 3.1.5 Easements An easement is an agreement by a landowner to give up certain property rights for some period of time – usually the right to change land uses and undertake other environmentally damaging activities – in exchange for a payment. If the title to the property is transferred, this restriction conveys with the title. We include easements in our discussion of PAs because they can be considered a form of private PA. Easements are popular in the United States, where they are a favored conservation tool of NGOs like Conservation International and The Nature Conservancy. However, easements are more problematic in Latin America, where property rights are sometimes complex, ill-defined, and politically contentious (Wolman 2004). Examples can be found in Mexico, where both local and foreign NGOs have cooperated to protect 40,000 acres in the Sierra Madre Occidental, and where the Worldwide Fund for Nature and the Mexican government are cooperating to purchase logging rights from local inhabitants in monarch butterfly habitat (Wolman 2004).

3.2 Forest comanagement 3.2.1 Description Forest comanagement refers to devolution of control over forests from central or regional governments to local communities. Although the institutional arrangements vary – they include indigenous territories, mixed-use PAs (i.e., extractive reserves), and communitymanaged concessions – all have several common features (Pacheco et al. 2012). First, the state formally allocates to local communities property rights to the resources. In most cases, these rights include access, harvesting, management, and the ability to exclude other users, but not alienation – that is, the right to sell, lease, or transfer the resources. Second, property rights are usually granted to communities or collectives instead of individuals. And finally, they are typically provided on the condition of “sustainable” use.



3.2.2 Status and trends A global movement toward forest comanagement began in the mid-1980s and became prominent a decade later (Agrawal et al. 2008). Drivers included fiscal and administrative constraints on continued state management, local communities’ demands for participation, external pressure from bilateral, multilateral, and other donors, and emerging scholarly work on local participation (Agrawal et al. 2008; Larson and Soto 2008). Although comprehensive, up-to-date statistics on the status and trends of forest comanagement in LAC are scarce, the information we do have clearly reflects a significant, if not momentous, increase. As of 2002, roughly 195 million hectares of forest and forest mosaics had been granted to local communities in 16 Latin American countries (White and Martin 2002; Pacheco et al. 2012). This land represents 22 percent of all forestland in the region (FAO 2011). Although there is about twice as much land in terrestrial PAs in LAC, PAs include both forests and other types of land cover (IUCN 2011). The countries leading the comanagement charge are: Brazil, where 100 million hectares has been granted to about half a million people; Bolivia, where 20 million hectares has been granted to indigenous groups; Nicaragua, where three million hectares has been granted to indigenous groups; and Guatemala, where about half a million hectares has been granted as eco-certified community forest concessions (Pacheco et al. 2012). The movement toward forest comanagement in LAC, and particularly in Central America, is reflected in FAO forest tenure data (Table 3.2-1; for more detailed country-level data see Tables A3.2-1 and A3.2-S1). Worldwide, 80 percent of forests are publicly owned and 18 percent are privately owned, including by local communities. In Central America, by contrast, only 52 percent of forests are publicly owned and 46 percent are privately owned. Of Central America’s privately owned forests, 40 percent are owned by communities (FAO 2010). The movement toward forest comanagement in LAC also is apparent from 2008 statistics on tenure in the 39 countries with the most tropical forest (Table 3.2-2).1 Among the LAC countries in this group, 32 percent of forests are either owned or managed by local communities, versus 27 percent in Asia and less than 1 percent in Africa. 3.2.3 Issues Theoretical and empirical research suggests comanagement can either stem or spur forest damage. Stemming forest loss In LAC, de facto property rights on state-managed forests are often ill-defined owing to governments’ limited ability and willingness to monitor and enforce restrictions on use and exploitation. Table 3.2-1 Forest tenure by subregion, 2005 Region




World Caribbean Central America South America

80 82 52 77

18 16 46 19

2 2 2 4

Source: FAO 2010.



Table 3.2-2 Forest tenure by subregion for 39 countries with most tropical forest, 2008 (%) Region

Public Designated for use by community Owned by community or Individual or or indigenous group indigenous groups firm

Latin America 36 Africa 98 Asia and Pacific 67

7 0 3

25 0 24

32 2 6

Source: Dahal et al. 2011.

It is well known that such situations create incentives for agents to overexploit natural resources – that is, they can contribute to a “tragedy of the commons” (Hardin 1968; Bromley 1992). Analyses of the tropical agricultural frontier find that ill-defined and insecure property rights can spur deforestation by encouraging land-poor households to colonize frontier areas (Clark 2000; Oliveira 2008); strengthening colonists’ preferences for unsustainable productive activities with quick returns versus investments in forests and other long-lived assets (Mendelsohn 1994; Barbier and Burgess 2001); inciting squatters to clear forest to establish use rights or block competing claims (Araujo et al. 2009; Damnyag et al. 2012); and preventing land managers from participating in conservation initiatives like payment for environmental services (PES) and reduced emissions from deforestation and degradation (REDD) schemes (Wunder 2005; Gregersen et al. 2010).2 In principle, comanagement could mitigate each of these problems. In addition, some authors have highlighted advantages of comanagement over PAs, the cornerstone of most countries’ conservation efforts (Molnar et al. 2004; Bray et al. 2008; Porter-Bolland et al. 2012). Although political and economic factors limit the amount of land that can be formally protected, comanagement does not face that constraint. Also, in principle, comanagement mitigates trade-offs between conservation and economic development by allowing for sustainable use of resources. Spurring forest loss Unfortunately, theoretical and empirical research also suggests that comanagement can spur deforestation, particularly when it entails granting property rights to communities, which as noted above, is the norm. Community comanagement can recreate common-pool resource problems on a local level, and communities vary considerably in their ability and willingness to address these problems (Ostrom 1990; Persha et al. 2011). Moreover, community control over forests can be undermined or coopted by powerful private actors or “reappropriated” by central governments (Johnson and Forsyth 2002; Engel and López 2008; Ribot et al. 2006; Barrett et al. 2001). 3.2.4 Effectiveness Given that comanagement can either stem or spur unsustainable resource use, the net effect is an empirical matter. But evaluating the causal effect of comanagement on resource use is challenging because it is not randomly assigned across communities. Rather, state authorities tend to implement comanagement in communities with above-average preexisting management capacity (Bowler et al. 2012). Evaluations that fail to control for this tendency will conflate the effects of comanagement on outcome measures such as forest cover change with the effects of preexisting characteristics of participating, thereby generating overly optimistic assessments.



Unfortunately, the evidence base on the causal effects of comanagement on resource use is quite thin, and many studies are plagued by failure to control for selection bias and other problems. Bowler et al. (2012) present a comprehensive review of the literature on the causal effects of forest comanagement on environmental and socioeconomic outcomes. They find that, despite considerable anecdotal evidence, few rigorous studies on the topic have been undertaken. Their search identified only 42 studies, more than three-quarters of which had significant methodological problems. They conclude that there is not enough evidence to generalize about whether and under what conditions comanagement is an effective conservation policy. A complicating factor is that, as noted above, comanagement includes very different institutional arrangements – ranging from certified community forest concessions to indigenous territories that entail land-use restrictions – that almost certainly have very different effects on biodiversity conservation. Based on their own review of rigorous studies, Miteva et al. (2012) reach similar conclusions. In any case, those broad findings jibe with general conclusions of complementary efforts to identify the drivers of the effectiveness of comanagement of forests and fisheries. These studies generally find that place-based institutional and contextual factors determine effectiveness, including leadership, social capital, careful definition of property rights, participation by resource users in management, accountability of decision makers, and effective monitoring of outcomes (Agrawal et al. 2008; Pagdee et al. 2006; Evans et al. 2011; Gutiérrez et al. 2011). Given the lack of evidence about the effectiveness of comanagement as a conservation strategy, along with emerging consensus that place-based factors drive effectiveness, broad policy recommendations about using comanagement to further biodiversity conservation are risky. However, several authors have argued that, as in the case of PAs in LAC, comanagement already is a very important component of the policy landscape – as noted above, more than one-fifth of LAC forests are already under some form of comanagement – and therefore, it is imperative that governments, donors, and other stakeholders devote the resources needed to ensure that comanagement is as effective a conservation tool as possible (Pacheco et al. 2012; Sunderlin et al. 2008; Molnar et al. 2004; Barrett et al. 2001). For example, Pacheco et al. (2012) argue that forest comanagement is in some ways analogous to agrarian reform, which has been a centerpiece of LAC rural development policy. But whereas considerable resources – including infrastructure, subsidies, technical assistance, and research – were devoted to ensuring that agrarian reform would succeed, such support has yet to be devoted to supporting comanagement.

3.3 Land-use planning 3.3.1 Description The objective of land-use planning is to identify and promote the adoption of land uses that best meet the needs of the local population while safeguarding future resource availability (FAO 1993). Hence, it must balance competing demands for land among industry, agriculture, tourism, housing, transportation, and – most relevant here – biodiversity conservation and ecosystem service provision (McNeil et al. 2012). Land-use planning that incorporates the last two considerations can be used to further conservation goals by restricting agricultural expansion into ecologically sensitive areas and/or confining it to already-degraded lands; protecting watersheds; establishing corridors to maintain biodiversity; controlling ecological damage from coastal development; and enabling marshes, mangroves, and lagoons to migrate inland with sea-level rise.



Land-use plans can be indicative or prescriptive. Indicative plans are used to select land uses on particular sites. For example, they are used by governments to site new roads, mining operations, logging concessions, and protected areas. Prescriptive plans, by contrast, outline what land uses are allowed or disallowed on a larger landscape and are often used to address environmental externalities (Chomitz 2007). Although land-use planning can occur at the national, district, and local levels, the last two levels are particularly important (FAO 1993). Whereas national land-use policy may entail laws and regulations governing land tenure, forest clearance, and water rights, the actual siting of new developments and infrastructure typically occurs at the district level (FAO 1993). More localized planning, such as at the village level, tends to involve a particularly bottom-up approach (FAO 1993). Several broad approaches to land-use planning have been used. Within agricultural and forestry science, information on soil, topography, crop models, people, and forest cover are used to develop recommended land uses across a region. This approach is common in Bolivia, Peru, Chile, Colombia, Costa Rica, Nicaragua, and Paraguay (Chomitz 2007). A second broad approach, which focuses on environmental protection, is to arrange land uses to meet environmental goals at the least cost (Margules and Pressey 2000; Chomitz 2007). Sometimes referred to as conservation return on investment analysis, it has been used to propose plans for siting protected areas (Boyd et al. 2012). A third land-use planning strategy is the ecosystem approach, which was adopted by the Convention on Biological Diversity in 2000 as the main framework for achieving objectives of conservation, sustainable use, and equitable distribution of biodiversity goods and services (Andrade Perez 2008). It emphasizes adaptive management and consideration of indigenous and local knowledge and culture, employs a long-term perspective, and considers the role of healthy ecosystems in the production of goods and services. It has been applied in a variety of contexts within LAC, especially in the context of protected area planning. Finally, ecological-economic zoning (EEZ) is a planning approach that emphasizes the interaction between biophysical and socioeconomic factors (FAO 1997). It seeks to identify optimal land-use patterns (to be enforced through institutional action) that reduce or mitigate negative environmental externalities, a focus that makes it a particularly useful approach in ecologically sensitive regions (Nogueira et al. 2000). Mostly used in large land areas like river basins with a substantial human population, EEZ requires spatial data. 3.3.2 Status and trends Buenos Aires’ 1977 land-use planning law initiated formal land-use planning in LAC. In the next two decades similar laws were passed in Colombia, Chile, and other countries (Cabeza 2002). However, land-use planning did not incorporate environmental considerations until the 1990s, when Bolivia, Ecuador, El Salvador, Honduras, Nicaragua, and the Dominican Republic began to employ land-use zoning as an instrument for environmental protection. Since then, increased consideration of environmental goals in land-use planning has been a major trend in the region. For example, leaders in LAC have explicitly recognized the importance of implementing coordinated land-use planning systems that will improve management of the environment and natural resources (Montes Lira 2001). Montes Lira (2001) presents case studies of urban land-use planning in three LAC cities: Montevideo, San Salvador, and Bogotá. In Montevideo, population growth has been moderate, and urban expansion into agricultural land is an emerging issue. The environmentally focused objectives of Montevideo’s land-use plan are limiting unnecessary urban expansion,



increasing the density of areas with existing infrastructure, preserving productive land and protecting ecosystems, and coordinating land-use plans with city plans for sanitation and other infrastructure. Although Montevideo has a legal framework for developing a plan, it lacks some of the regulatory tools needed for its implementation. San Salvador, on the other hand, is the main destination for rural–urban immigrants in El Salvador, and has experienced tremendous unplanned growth and gaps in transportation and sanitation infrastructure. The city has developed an integrated land-use and environmental plan for the metropolitan area, with special consideration for seismic, hydrological, and sanitary issues and regulations. Major challenges have been the lack of jurisdiction over other municipalities within the San Salvador metropolitan area and a weak legal framework for implementation. In Bogotá, as in San Salvador, population growth has pushed the boundaries of the city into agricultural lands where infrastructure is lacking. The main environmental objectives of Bogotá’s land-use plan are improving management of natural resources, maintaining agricultural productivity, and organizing land-use according to comparative advantages of different locations. Although Bogotá has the legal framework needed to develop and implement a plan, the process faces challenges, especially in coordination with other municipalities (Montes Lira 2001). Nonetheless, Bogotá’s efforts have been considered successful, in part because they have encouraged participation by citizens and businesses and created local institutions and information systems that have allowed various stakeholders to monitor implementation (UNEP 2010). In Mexico, a legal framework exists for municipalities to implement land-use planning. However, states have issued many ordinances despite lacking clear jurisdiction to do so, and indigenous tribes have implemented land-use plans within their own communities inside municipalities. The various land-use plans create jurisdictional problems and call for improved coordination (UNEP 2010). Finally, EEZ has been used extensively in Brazil, especially in the Amazon region. It has been present in policy and planning documents since the 1980s and became integral to national policy planning by the 1990s (Nogueira et al. 2000). By 1991 the state of Rondonia had a “first approximation” of a land-use plan that divided the state into six macro zones. Zones 4 through 6, which constituted 55 percent of the state, were declared off-limits to most economic activity, and Zone 1 was broken into seven subzones that specified which crops were appropriate in each area (Mahar and Ducrot 1998). EEZ has become an important policy lever for conservation in Brazil, and it often interacts with the agroecological zoning, which delimits sugar cane farming to prevent expansion into important conservation areas. In 2010 an Amazon ecological-economic macrozoning law was approved, providing coarse guidance for land use in the region while meeting conservation goals (Government of Brazil 2010). 3.3.3 Issues As the above case studies make clear, weak governance, particularly at the local level, is a major challenge to effective land-use planning (Montes Lira 2001; Cabeza 2002; FAO 1993). Specifically, local and state regulatory institutions implementing land-use planning lack technical capacity, laws, and regulations to ensure implementation, and perhaps most important, coordination across levels of government and government sectors (e.g., transportation, infrastructure, and agriculture). Additional challenges have to do with inherent uncertainty, enforcement, property rights, distributional issues, and participation (Chomitz 2007; Montes Lira 2001). Land-use planning requires anticipating the future effects of land uses on people



and the environment. However, these effects depend on complex general equilibrium and ecological dynamics and therefore are highly uncertain. At the same time, short planning horizons driven by political considerations, among other factors, have led to land-use plans that focus mainly on the near future (Cabeza 2002). Land-use zoning, by design, must restrict the rights of current and future landowners. As a result, enforcement is critical, as are land tenure issues. Distributional issues arise because land-use planning can impose burdens on the poor, who are less able than the rich to circumvent and adapt to land-use planning restrictions. Finally, participation by local stakeholders can greatly improve land-use planning but is often not facilitated in LAC governance.

3.4 Fisheries management 3.4.1 Description LAC’s diverse coastal and marine environments endow the region with some of the most productive fisheries in the world (Salas et al. 2011). Fisheries generate considerable economic activity, ranging from small-scale and artisanal fisheries focused on domestic markets to large commercial operations oriented to international markets. Coastal populations, especially small-scale fishers, often heavily depend on fisheries for income and food. 3.4.2 Status and trends Economic importance Fishing is an important economic activity in LAC. In 2004 the share of fisheries in GDP was 6 percent in Ecuador, 5 percent in Belize, 4 percent in Colombia, 3 percent in Chile, and at least 2 percent each in the Bahamas, Grenada, Guyana, Panama, Peru, and St. Vincent and the Grenadines. Fisheries in Chile, Mexico, Colombia, and Brazil each contributed more than $2 billion to national GDP; those in Venezuela, Panama, Argentina, Guyana, and Peru contributed more than $100 million (Bovarnick and Alpízar 2010; see Tables A3.4-S1 and A3.4-S2 for country-level data). According to FAO (2010), 1.3 million people work as fishers or fish farmers in LAC. Roughly two-thirds of the total is in capture fisheries, and the rest in fish farming. Whereas in much of the developed world fisheries-related employment is declining even as the total output grows, fishing employment in LAC has continued to increase in the past two decades, albeit at a moderate pace (FAO 2010). Fleet and market characteristics The fishing sector in LAC is dominated by small vessels (less than 12 meters), which account for nearly 95 percent of the region’s collective fishing fleet (FAO and World Fish Center 2008). Much of the fish and fisheries products from LAC are consumed domestically, but exports are important and growing. LAC as a whole is a net exporter of fish and fish products – imports to the region total less than $4 billion, and exports are $12 billion. World demand for fishery exports has steadily increased and outpaced population growth, a trend that is expected to continue (FAO 2010). Within LAC, Peru accounts for the majority of total catch. Chile, Argentina, Mexico, Brazil, and Venezuela are large contributors. Although fish landings are relatively smaller in



the Caribbean, they provide an important source of export income to some countries (Salas et al. 2011). Production in capture fisheries As discussed in Section, production in LAC capture fisheries, particularly Caribbean ones, has already plateaued and is declining. 3.4.3 Issues Rebuilding depleted fisheries and restoring essential fish habitat require a comprehensive approach to management, governance, and regulations; science, information systems, capacity building, and public awareness are necessary components (Gelcich et al. 2010; Hilborn et al. 2003; Worm et al. 2009; World Bank 2009; Salas et al. 2011). Heterogeneity of environmental, institutional, and economic conditions within the region calls for different solutions in different locations (Cinner et al. 2012). Moreover, disparities between industrial and artisanal fisheries imply they should be treated differently in fisheries management and policy (Defeo and Castilla 2005). Even though specific solutions to promote sustainable fisheries vary by location and fishery, a consensus on the broad outlines of the best policy solutions is emerging among experts. The following recommendations draw and expand on the findings from a 2003 FAO workshop on fisheries (Swan and Gréboval 2004). •

Grant fishers individual or collective rights to the fishery. An increasingly common approach to managing fisheries is to grant fishers rights. These rights – sometimes referred to as catch shares or dedicated access privileges – can be to a certain number of fish or to the exclusive use of a defined area. They can be granted to individuals or cooperatives, they can be tradable or not, and the total allowable catch can be determined by the right holders or a regulator. Properly implemented, the approach can substantially increase economic benefits from fisheries. The guiding rationale is the same one that underpins forest comanagement implemented via tenure reform: giving harvesters a secure asset creates incentives to manage it sustainably. In an article published in Science, Costello et al. (2008) studied rights-based fisheries management using panel data on more than 11,000 fisheries and found that “the implementation of catch shares halts, and even reverses, the global trend toward widespread collapse” and “has the potential for greatly altering the future of global fisheries.” Use market and economic incentives to address overharvesting and economic inefficiencies. Options include individual transferrable quotas (ITQs), which can more efficiently allocate catch-shares across fishers and ensure the most efficient fishers do most of the fishing (e.g., Copes 1986; Sanchirico and Newell 2003), and reform (reduction, elimination, or decoupling) of harmful subsidies that encourage the overbuilding of fishing capacity (see Section 3.7). In addition, economic tools can be used to distribute benefits from the fishery in ways that better address the needs of local communities. Conserve critical coastal and marine areas. Increased conservation of coastal and marine areas that are particularly important ecologically is needed to ensure sufficient nursery habitat, protect overall biodiversity, and limit catch. Marine PAs are increasingly popular for improving the preservation of critical coastal and marine habitats and



their biodiversity (Kelleher 1999; Guarderas et al. 2008; see also Section 3.1 for more information on PAs). They now cover altogether 11 percent of territorial waters (up to 12 nautical miles from the shore (Section 3.1, Table 3.1-1). The increase in the size of protected areas has been roughly fourfold since 1990. However, current protected areas are unevenly distributed within LAC. Although a few ecoregions and countries (e.g., Venezuela) have large coverage, there are also many ecoregions and areas with only minimal protections. For example, the Pacific coast is underrepresented in protected areas (Chatwin 2007). Moreover, in the Caribbean, despite its threatened marine and coastal biodiversity, only about 2 percent of the marine area is protected (IUCN 2011). Evaluating options for new marine PAs requires case-by-case assessment because marine ecosystems are complex and highly connected systems. Moreover, PAs cannot be evaluated in isolation, and their effectiveness can critically depend on spillovers and fisheries management outside protected areas (Gaines et al. 2010; White et al. 2008). For example, while improved siting of marine PAs can substantially increase the economic benefits they generate, these benefits can be greater still when fisheries management is fully optimized both spatially and economically (Rassweiler et al. 2012). Moreover, extending the size of marine reserves may be less advantageous than networks of smaller protected areas (Halpern et al. 2009). PAs aside, other alternatives to promote conservation of coastal and marine areas include more ecologically oriented coastal planning, such as reduced coastal development to avoid habitat loss and degradation and coastal pollution, and the use of ecological offsets to mitigate habitat loss and degradation induced by development. Promote transparent and participatory management. Providing a meaningful role to different stakeholders in a range of fisheries management tasks and decisions can help improve the sustainability of fisheries. Management transparency promotes trustworthiness and can help reduce all forms of corruption. Stakeholders’ participation helps develop a sense of ownership and responsibility over the resource (Swan and Gréboval 2004). Support science, enforcement, and planning. Fisheries management is seriously limited by the lack of scientific knowledge, capacity, and data. Supporting all aspects of knowledge creation and sharing can therefore help improve fisheries management. For example, our limited information on the status of fish stock resources hinders or altogether prevents improved fishery management, such as the use of catch shares. Lack of information on the fluctuations of fishery stocks due to natural causes (such as El Niño and La Niña) thwarts efforts to adapt fisheries management accordingly (Swan and Gréboval 2004). Integrate policy and planning. Fisheries are situated in the broader coastal and marine ecosystems, which are dynamic, integrated, and subject to a broad range of interacting ecological and economic factors. As a result, fisheries policy and planning can be improved by adapting more integrated and ecosystem-based approaches to coastal and fisheries management (e.g., Halpern et al. 2010). Build capacity and public awareness. Improving and reforming fisheries policy and management requires considerable political and economic efforts, which are often thwarted by a lack of political will. Therefore, efforts to better inform policy makers and the general public about fisheries issues, including their economic and ecological importance, can help promote sustainability. Capacity-building efforts can benefit from improved science and may focus on promoting specific fisheries management instruments, such as comanagement or market-based instruments (Swan and Gréboval 2004).


Policies Comanagement in combination with marine PAs also holds promise. It can improve governance and fisheries performance and reduce management costs. Initiatives to support and evaluate fisheries comanagement could significantly improve fisheries management.

Table 3.4-1 summarizes the main factors of unsustainability of fisheries and potential solutions to address them. 3.4.4 Comanagement Comanagement of fisheries, the coastal and marine analog of forest comanagement (discussed in Section 3.2), has attracted increasing attention in recent years. Advocates assert that cooperative management by fishers, other local community members, fisheries managers, and scientists will promote sustainable resource use. Case studies of comanagement are now available, although most have methodological problems. Gutierrez et al. (2010) report on a systematic evaluation of 130 comanaged fisheries across the world, including many in LAC. The results echo those from meta-analyses of forest comanagement: sustainability does not proceed from comanagement alone but depends as well on place-based factors. Strong leadership turns out to be the most important factor, followed by individual and community quotas, strong social cohesion, and the presence of PAs. Less important criteria include enforcement and the existence of a long-term management policy. Importantly, successful comanagement generally requires the simultaneous presence of several contributing factors. In another systematic study, Cinner et al. (2012) evaluate more than 40 comanaged fisheries across five countries, arguing that comanagement is by and large successful in achieving its social and ecological goals. But the study also finds that comanagement tends to benefit wealthier resource users and that overexploitation is most strongly influenced by market access and the dependence of users on the resources. The study identifies institutional characteristics as having a strong influence on livelihood and compliance outcomes, but little effect on ecological conditions.

Table 3.4-1 Challenges and solutions to unsustainability of fisheries

1 2 3 4 5 6

Main factors of unsustainability

Potential solutions [main factor of unsustainability addressed]

Inappropriate economic incentives and subsidies High demand for limited resources

Improve rights (catch-shares, comanagement, etc.) [1, 2, 5] Enhance or create market and economic incentives (ITQs, PES, subsidy reform, etc.) [1, 2, 3, 5, 6] Preserve coastal and marine areas [2, 4, 5, 6] Transparent and participatory management [5] Support science, enforcement, planning [4, 6] Integrated policy and planning [4, 5, 6] Capacity building and public awareness building [1, 2, 5]

Poverty and lack of alternatives Fisheries’ complexities, lack of knowledge, natural fluctuations Lack of governance Interactions among fishery and other sectors, including environmental impacts

Source: adapted and augmented from Swan and Gréboval 2004.



3.5 Wastewater treatment 3.5.1 Description The main contributors to water pollution in LAC are agricultural production, which is a nonpoint source, and households and industry, which are point sources. In most LAC countries, nonpoint sources are more important than point sources. For example, in the early 1990s, a study of organic water pollution in Colombia’s surface waters estimated that agriculture contributed 84 percent, households 10 percent, and industry 6 percent (Blackman et al. 2006). Among point sources, households are generally far more important than industry. For example, in Colombia in the early 1990s, roughly three-quarters of organic pollution was generated by households and one-quarter by industry (Blackman et al. 2006). Nonpoint source pollution is a difficult problem because contributors are, by definition, small and dispersed. Strategies for addressing it entail preventing (rather than treating) pollution by reducing application of fertilizers, reusing irrigation water, and using vegetative barriers and other means to stem runoff (Ongley 1996). Agricultural nonpoint sources are discussed in Section 3.15. In this section, we focus on point sources. Point source wastewater can be treated to remove organic and inorganic pollutants. Treatment is usually categorized as preliminary, which involves screening and other simple processes; primary, which entails sedimentation (i.e., letting wastewater stand so that pollutants settle out); secondary, which involves using biological processes to break down pollutants; and tertiary, which entails removing pollutants such as nitrogen and phosphorus (WEHLL 1999). Treatment can be either centralized in large-scale plants or dispersed in small units, such as septic tanks. In industrialized countries, the most common form of centralized secondary treatment is called the activated sludge process, an energy-intensive, sophisticated approach that requires careful operation of heavy machinery. In developing countries, less sophisticated methods are sometimes used, including stabilization ponds that rely on sunlight and algal growth to break down pollutants; constructed wetlands, which are simple holding ponds filled with aquatic plants; trickling filters, which run wastewater over loose beds of stones; and upflow anaerobic sludge blankets, which use a layer of bacteria to absorb pollutants (Noyola et al. 2012; WEHLL 1999). 3.5.2 Status and trends In LAC, the vast majority of wastewater is not treated in any way. To our knowledge, the most recent, reasonably comprehensive region-wide data are from 2000 (PAHO 2001). In that year, 86 percent of LAC wastewater was not treated. Treatment rates varied from a low of 0 percent in several Caribbean countries to a high of 100 percent in several other Caribbean countries (Table 3.5-1). Among the region’s most populous countries, Mexico treated only 15 percent of its wastewater, Colombia 11 percent, and Brazil 10 percent. Wastewater treatment is often confined to selected cities. For example, in 1999, only 16 percent of Colombia’s 1,089 municipalities had operating wastewater treatment plants (Blackman et al. 2006). A study of secondary wastewater treatment in representative LAC countries found that the most common types are stabilization ponds, activated sludge, upflow anaerobic sludge blankets, and trickle filters (sometimes used in combination), which together represent 80 percent of all wastewater treatment in the region (Noyola et al. 2012).



Table 3.5-1 Percentage of sewer water with some type of treatment, by country, 2000 Americas Country Argentina Belize Bolivia Brazil Chile Colombia Costa Rica Cuba Ecuador El Salvador Guatemala Honduras México Nicaragua Panama Paraguay Peru Uruguay Venezuela

Canada United States

Caribbean Percentage 10 57 30 10 17 11 4 9 5 2 1 3 15 34 18 8 14 77 10

80 100



Anguila Antigua and Barbuda Aruba Bahamas Barbados Bermuda British Virgin Islands Cayman Islands Dominica Dominican Republic Dutch Antilles French Guyana Grenada Guadalupe Guyana Haiti Jamaica Martinique Montserrat Puerto Rico San Kitts and Nevis San Vicente and Grenadines Santa Lucia Suriname Trinidad and Tobago Turks and Caicos Islands US Virgin Islands

NA 100 ND 80 100 ND 0 ND 0 49 ND 65 0 40 50 0 ND ND 100 100 ND ND 46 0 65 0 ND

Source: PAHO 2001.

3.5.3 Issues Historically, a large share of wastewater treatment in LAC has relied on activated sludge and other sophisticated, energy-intensive methods used in industrialized countries. Such facilities are expensive to build, maintain, and operate. As a result, few wastewater treatment plants have been built, and most of those are inefficiently and only intermittently run (WHELL 1999). For example, in Colombia in the late 1990s, almost one-fifth of wastewater flows into treatment plants were not actually treated in any way, and 40 percent of plants failed to treat wastewaters to minimum regulatory standards (Blackman et al. 2006). As noted above, a variety of less expensive and less complicated – but still effective – treatment options are available, including stabilization ponds and constructed wetlands. In addition, two alternative treatment strategies have been advocated: dispersed treatment, which entails lower fixed costs than centralized treatment and diversifies risks of equipment and human failures; and the recovery and reuse of wastewaters (typically for irrigation), which requires treating wastewater to a relatively low standard (Evans 2013; Wade 2012; WHELL 1999).



3.6 Environmental governance 3.6.1 Description Governance is the exercise of economic, political, and administrative authority to manage a country’s affairs. Good governance is needed to both develop and enforce regulatory approaches to environmental and natural resource management, including both mandates and incentive-based policies. Efforts to improve governance generally focus on increasing the voice of various stakeholders, accountability, public service delivery, transparency, security, and political stability. 3.6.2 Status and trends Environmental laws, policies, and programs Most countries in LAC have established environmental laws, policies, and programs (UNEP 2010b). For example, 12 of 14 countries in South America have specific forest laws, eight have forest policies, and ten have programs; and six of seven countries in Central America have laws, policies, and programs (Table 3.6-1; see Table A3.6-1 for country-level data). Less information is available for the Caribbean. However, the existence of laws, policies, and programs does not necessarily mean they are effective. As discussed below, in LAC, as in many developing countries, environmental laws and policies are often not monitored or enforced, and programs not implemented. Governance effectiveness Governance is often evaluated using the World Bank’s Worldwide Governance Index (WGI) comprising indicators of voice and accountability, political stability, government effectiveness, regulatory quality, rule of law, and control of corruption (World Bank 2012). Of these indicators, three are particularly relevant for evaluating environmental governance: (1) government effectiveness, which measures the quality of public services, the capacity of the civil service and its independence from political pressures, and the quality of policy formulation; (2) regulatory quality, which measures the ability of the government to provide sound policies and regulations that enable and promote private sector development; and (3) rule of law, which measures the extent to which agents have confidence in and abide by the rules of society, including the quality of contract enforcement and property rights, the police, and the courts, as well as the likelihood of crime and violence. Table 3.6-1 National forest policies, programs, and laws in LAC, by subregion Subregion

South America Central America Caribbean Source: FAO 2011.





No No Yes No No Specific Incorporated No data data forest law in other law law

No data

8 6 10

6 0 4

0 1 12

0 1 13

10 6 8

4 0 6

0 1 13

12 6 10

1 0 3

1 0 2



On average, LAC countries have scores on these indices that are midway between weak and strong (Table 3.6-2). Within the region, South America and Central America tend to have weaker governance scores, on average, than the Caribbean. Based on these indices, Venezuela has the weakest governance in South America and Chile has the strongest, and Costa Rica has the strongest governance in Central America. Also, the Caribbean and South America have greater variation in governance scores than Central America. As in most developing countries, the main limitation of environmental and natural resource management in LAC is weak enforcement (Chomitz 2007; UNEP 2010b; Russell and Vaughan 2003; Eskeland and Jimenez 1992). Regional and in most cases national data on monitoring and enforcement are not available. However, case studies of environmental regulatory monitoring and enforcement in countries like Colombia, reputed to have a relatively strong environmental management regime, make clear that grossly inadequate monitoring and enforcement are ubiquitous (Blackman and Sisto 2006; Blackman 2009; Blackman and Kildegaard 2010; Caffera 2007; Escobar and Chávez 2011). For example, a case study found that command-andcontrol water pollution regulation in Colombia c.2000 was undermined by four weaknesses (Blackman 2009). First, 40 percent of the 32 regional environmental authorities in charge of monitoring and enforcing such regulation did not have complete, or even nearly complete, lists of major polluters. Second, although all water polluters were supposed to have discharge permits, less than one-third did. Third, very few plants were inspected. For example, in Bogotá, environmental authorities set a goal of inspecting just 30 percent of permitted polluters. Finally, 40 percent of regional authorities lacked technical equipment to monitor compliance. Weak monitoring and enforcement are pervasive in LAC natural resource governance as well. For example, in the 1990s, 80–90 percent of forest clearing in both Brazil and Bolivia was illegal (Tacconi et al. 2003). Lack of governance along with ill-defined property rights can be a particular problem far from population centers, resulting in illegal logging and clearing in areas that are often rich in biodiversity (Clark 2000; Oliveira 2008).

Table 3.6-2 Worldwide Governance Index indicators for LAC (–2.5 to +2.5), 2010 Region LAC Mean S.D. Caribbean Mean S.D. Max. Min. Mesoamerica Mean S.D. Max. Min. South America Mean S.D. Max. Min.

Government effectiveness 0.22 0.75

Regulatory quality 0.21 0.76

Rule of law 0.07 0.87

0.57 0.72 1.52 (Anguilla) –1.61 (Haiti)

0.48 0.70 1.37 (Anguilla) –1.57 (Cuba)

0.51 0.74 1.42 (Anguilla) –1.35 (Haiti)

–0.27 0.45 0.32 (Costa Rica) –0.96 (Nicaragua)

0.04 0.36 0.51 (Costa Rica) –0.47 (Belize)

–0.52 0.48 0.50 (Costa Rica) –1.04 (Guatemala)

–0.04 0.67 1.18 (Chile) –1.02 (Venezuela)

–0.14 0.86 1.44 (Chile) –1.58 (Venezuela)

–0.28 0.86 1.29 (Chile) –1.64 (Venezuela)

Source: World Bank 2012.



In LAC, as in most developing countries, underlying drivers of weak monitoring and enforcement are a lack of political will; limited financial and human resources; inadequate technologies needed to monitor compliance with environmental regulations; jurisdictional, procedural, and interagency inefficiencies; and a preponderance of hard-to-monitor, smallscale firms and farms (UNEP 2010b; Akella and Cannon 2004; Russell and Vaughan 2003; Eskeland and Jimenez 1992). Akella and Cannon (2004) use an economic framework to analyze how weak enforcement affects governance. They begin with the proposition that enforcement systems fail because enforcement does not generate a sufficiently large expected penalty from illegal behavior relative to expected profits. The expected penalty, in turn, depends on the likelihood and magnitude of the sanction. They develop location-specific estimates of the expected revenue and penalty from various illegal activities. They find that expected revenue from illegal logging in the Atlantic forest of Brazil is an order of magnitude greater than the expected penalty and that an even greater gap exists for illegal hunting and wildlife trade in Chiapas, Mexico. Those barriers are not insurmountable, however. Recently, monitoring and sanctioning of unlawful deforestation activities in the Brazilian Amazon have improved significantly. Federal police activity has increased, the Brazilian Environment and Renewable Resources Institute (IBAMA) has increased monitoring, and the National Monetary Council decreed that individuals and enterprises with a history of illegal actions affecting the environment would be denied credit access (UN 2010). According to Akella and Cannon (2004), improved enforcement requires a comprehensive approach, including strengthening of monitoring, enforcement, and judicial and regulatory improvements; focusing on only one of these determinants is unlikely to generate positive results. Environmental capacity and mainstreaming Even though environmental protection provides many public benefits, these benefits are not taken into account in most policy making, and hence conservation is underprovided. One reason is a lack of knowledge about environmental and natural resources issues – and the social, economic, and ecological benefits of conservation in particular – among both policy makers and their constituents (UN 2010). Again, the forest sector provides a useful illustration. In Central America, more than 60 percent of staff of forest agencies had a university degree. However, in South America fewer than half did, and in the Caribbean fewer than one-fifth did (Table 3.6-3). Table 3.6-3 Human resource levels in LAC, 2008 Region

South America Central America Caribbean

Information availability*

Staff, 2008

Countries Total forest area Total staff (percentage)

Per 100,000 ha

Percentage with university degree

12 6 11

1 6 115

47 61 13

93 54 54

5,215 1,167 4,146

Note *Number of countries and percentage forest area for which data are reported. Source: FAO 2011.



In addition, there are significant gaps in environmental capacity and training for those who manage landscapes, work in sectors that directly affect the environment, or decide on policy that affects the environment (Ceballos et al. 2009). The need for environmental capacity is growing as more management is decentralized, PAs are expanded, and the environment is increasingly pressured by industry, infrastructure projects, and other land-use changes. Although some countries (e.g., Peru, Bolivia) have sought to address this need by developing strategies for increasing and improving their environmental capacity, the strategies have not been fully implemented or funded (Ceballos et al. 2009). 3.6.3 Issues Although strategies for improving the effectiveness of environmental governance are necessarily place-based, an array of general strategies have been proposed. Among the most common is to increase decision makers’ awareness and understanding of the economic and social value of biodiversity and ecosystem services, including in nonenvironmental government activities, such as planning, finance, infrastructure, and economic development. In theory, such human capacity building will increase not only decision makers’ ability to design and implement effective policies, but also their willingness to do so; that is, it will enhance political will (UNEP 2010a; UN 2010). Human capacity in environmental management can be augmented by raising educational requirements for hiring, promotion, and professional certification of civil servants, targeted educational programs, and investments in reliable national statistics and scientific studies of environmental and natural resource issues (UN 2010). Beyond more human capacity, better enforcement is obviously needed for government mandates and incentive programs to be useful. Suggestions for enhancing enforcement in LAC include the following: • •

• • •

• •

strengthening mechanisms for stakeholders’ and citizens’ engagement to enhance political will and accountability; improving transparency by, for example, enhancing access to data on environmental performance (UNEP 2010b) and clarifying the process used to grant forest concessions (Chomitz 2007); improving the consistency and coordination of environmental policies and economic development policies by, for example, including environmental indicators in evaluation of projects and policies across economic sectors, so that economic considerations do not trump environmental concerns in land-use planning, environmental impact assessments, cost–benefit analyses, etc. (UNEP 2010b; UN 2010); increasing awareness of environmental regulations (UNEP 2010b); increasing credibility of enforcement institutions by, for example, raising the profile of ministries of the environment (UNEP 2010b; UN 2010); increasing human and technical capacity for monitoring environmental and natural resource conditions and ensuring compliance with environmental regulations, including by investing in geospatial technologies (UNEP 2010b; UN 2010); increasing financial resources devoted to enforcement (UNEP 2010b); and increasing the efficiency of regulations to minimize monitoring, enforcement, and compliance costs by, for example, using economic instruments (Chomitz 2007).

Another important component of environmental governance is the need for coordination at all levels, including among ministries, states, and countries. International coordination



is particularly important for the Caribbean, a region that is ecologically and environmentally interconnected but politically divided into many small countries and territories (UN 2010). B MARKET-BASED APPROACHES

3.7 Subsidy reform 3.7.1 Description Subsidies are government actions – monetary payments, transfers, or relief of opportunity costs – targeting a specific economic sector (Myers and Kent 2001). They can boost profits and therefore spur economic activity that damages biodiversity and ecosystem services. For example, biofuel price supports can encourage the expansion of cropped areas, and irrigation subsidies can contribute to overexploitation of aquifers. Hence, eliminating, reducing, or decoupling subsidies can benefit biodiversity and ecosystem services. Decoupling subsidies involves eliminating links to production levels and prices so that production decisions are not distorted by the subsidy. For example, the historical recipients of subsidies might receive lump-sum payments in lieu of payments tied to production. 3.7.2 Status and trends Agriculture Historically, LAC economic development policies, notably import substitution, tended to encourage industrialization and supported urban centers. As a result, the agricultural sector received relatively few government resources. Since the 1980s, however, many LAC governments have significantly increased support to agriculture to boost rural incomes and enhance food security. This support mainly aims at improving agricultural productivity (Anderson and Valdes 2008; Myers and Kent 2001). LAC governments subsidize agricultural outputs (with price supports), marketing, technical extension, and an array of inputs, including machinery, credit, fertilizers, pesticides, seeds, energy, and transportation. About half of LAC governments’ total rural expenditures are subsidies to goods like fertilizer and seeds that are privately owned by farmers, and the other half support public goods like transportation and irrigation infrastructure (Bulte et al. 2007). Although they pale in comparison with agricultural subsidies in many industrialized countries, LAC agricultural subsidies are significant and growing. For eight LAC countries that together account for 78 to 84 percent of LAC’s population, agricultural value added, and GDP (Argentina, Brazil, Chile, Colombia, the Dominican Republic, Ecuador, Mexico, and Nicaragua), on average the nominal rate of assistance – defined as the percentage by which government policies have raised (or lowered) gross returns to producers above (below) what they would be without the government’s intervention – rose from –7.2 percent in 1965–1969 to 4.8 percent in 2000–2004 (Tables 3.7-1–3.7-3). In constant 2008 dollars, for all countries in LAC, the gross subsidy rose from –$742 million per year to $5.4 billion per year, and per person engaged in agriculture, the average subsidy for all LAC countries rose from –$20 to $126 (Anderson and Valdes 2008). LAC biofuels subsidies have attracted considerable attention recently. Although Brazil’s ethanol program is probably the best known, many other LAC countries are planning or



Table 3.7-1 Nominal rate of assistance (NRA) in eight LAC countries, 1965–2004 (percent)* Country

1965– 1970– 1969 1974

1975– 1979

1980– 1984

1985– 1989

1990– 1994

1995– 1999

2000– 2004

Argentina Brazil Chile Colombia Dominican Republic Ecuador Mexico Nicaragua Weighted average

–22.7 –6.1 16.2 –4.7 5 –9.6 0 0 –7.2

–20.4 –23.3 4.5 –13 –21.2 –15 0 0 –18

–19.3 –25.7 7.2 5 –30.7 5.9 3.8 0 –12.5

–15.8 –21.1 13 0.2 –36.4 –1 3 0 –10.9

–7 –11.3 7.9 8.2 –1 –5.3 30.8 –3.2 4.2

–4 8 8.2 13.2 9.2 –2 4.2 –11.3 5.5

–14.9 4.1 5.8 25.9 2.5 10.1 11.6 –4.2 4.8

–22.9 –27.3 12 –14.8 –18.1 –22.4 0 0 –21

Note *NRA is the percentage by which government policies have raised gross returns to producers above what they would be without the government’s intervention (or lowered them, if NRA < 0). Source: Anderson and Valdes 2008.

Table 3.7-2 Gross subsidy equivalents of assistance to farmers in eight LAC countries, 1965–2004 (2008 $ millions) Country

1965– 1969

1970– 1974

1975– 1979

1980– 1984

1985– 1989

1990– 1994

1995– 1999

2000– 2004

Argentina Brazil Chile Colombia Dominican Republic Ecuador Mexico Nicaragua Eight-country total Region

–406 –189 114 –87 14 –47 0 0 –601 –742

–815 –2,531 108 –483 –145 –146 0 0 –4,012 –4,954

–996 –3,393 777 –712 –238 –187 0 0 –5,639 –6,962

–1,777 –7,700 163 378 –431 80 834 0 –8,454 –10,437

–1,132 –6,778 286 –7 –412 –22 539 0 –7,525 –9,290

–612 –2,991 332 802 –15 –111 6,418 –28 3,797 4,688

–569 2,968 443 1,488 142 –67 995 –133 5,267 6,503

–2,609 1,576 303 1,906 37 337 2,861 –57 4,354 5,376

Source: Anderson and Valdes 2008.

Table 3.7-3 Per person gross subsidy equivalents of assistance to farmers in eight LAC countries, 1965–2004 (2008 $) Country

1965– 1969

1970– 1974

1975– 1979

1980– 1984

1985– 1989

1990– 1994

1995– 1999

2000– 2004

Argentina Brazil Chile Colombia Dominican Republic Ecuador Mexico Nicaragua Eight-country total Region

–261 –12 154 –29 20 –49 0 0 –21 –20

–550 –154 147 –150 –203 –145 0 0 –130 –123

–698 –198 99 –200 –339 –184 0 0 –173 –165

–1,265 –445 198 99 –623 76 102 0 –251 –238

–778 –416 321 –2 –589 –19 64 0 –227 –211

–414 –201 350 216 –22 –91 749 –71 119 108

–387 214 456 399 225 –54 116 –334 170 150

–1,786 123 308 515 63 270 336 –144 147 126

Source: Anderson and Valdes 2008.



Table 3.7-4 Spending on irrigation in 13 LAC countries, 2001 Country

1995 PPP adjusted $

As percentage of total rural spending

Brazil Chile Costa Rica Ecuador Guatemala Honduras Mexico Nicaragua Peru Dominican Republic Venezuela Total

360.0 82.4 2.8 103.5 18.4 2.7 1,232.1 0.5 247.3 148.3 38.1 2,236.8

5.4 6.9 0.9 14.8 1.7 0.8 12.6 0.1 33.0 11.6 8.7

Note PPP = purchasing power parity. Source: Kerrigan 2007.

already have biofuel programs. Countries that have received IDB assistance for biofuels programs include Honduras, El Salvador, Chile, and Colombia (Vieira de Carvalho 2011). Some quick statistics from Mexican conservation and agricultural budgets provide an indication of how important subsidy reform could be to biodiversity conservation. In 2009 the budget of Mexico’s agricultural assistance program, PROCAMPO, was more than twice that of the national parks agency, CONANP, and Mexico’s assistance program specifically targeting livestock production, PROGAN, had a budget twice the size of PROCAMPO (CONANP 2012; PROGAN 2012; PROCAMPO 2012). Water Many LAC countries subsidize water used for human consumption and irrigation, often heavily. Subsidies contribute to overexploitation of surface water and groundwater, which in turn harms biodiversity and ecosystems by disrupting river ecosystems and damaging wetlands. Irrigation subsidies in particular tend to be substantial (Myers and Kent 2001). In 2001, spending on irrigation in 13 LAC countries ranged from $0.5 million in Nicaragua to $1.2 billion in Mexico. As a percentage of total rural expenditures, it ranged from 0.8 percent in Honduras to 33 percent in Peru (Table 3.7-4). As just noted, at the country level, Mexico was by far the biggest spender on irrigation subsidies, having devoted more than $1.2 billion to irrigation in 2001. In that year, farmers paid water prices that amounted to just 3–8 percent of production costs (Myers and Kent 2001). Electricity subsidies that reduced the cost of pumping water reinforced those created by irrigation water subsidies. Large-scale farms – the main beneficiaries of water and electricity subsidies – tend to use the most inefficient traditional irrigation methods (Guevara-Sangines 2006). Fishing Subsidies contribute to overexploitation of marine and freshwater fisheries in LAC (Schrank 2003). The exact inputs and outputs subsidized vary from country to country, but the most common subsidy is on fuel used by fishing vessels. Others include subsidies on vessels and



other capital equipment and variable inputs like bait and ice (FAO 2011). Awareness of the effect of fishing subsidies on overexploitation has been growing since a 1993 FAO study reported that the fishing industry was operating at global costs that exceeded commercial revenues by more than $50 billion per year (Virdin 2001). In 2003, LAC fishing subsidies totaled almost $1.5 billion (Sumaila et al. 2010). Brazil’s subsidies, which accounted for the largest share, totaled more than $400 million. Other LAC countries with large total fishing subsidies are Argentina ($366 million) and Peru ($205 million). Although LAC total subsidies are lower than those of many regions, along with Africa’s they are the highest as a share of landed value, which for LAC is close to 50 percent. Energy Perhaps the most common type of subsidy in LAC – and in other developing and industrialized regions – is to energy, specifically to fossil fuels and electricity. Energy subsidies are popular because they supposedly have a progressive distributional effect and boost international competitiveness. A recent report examines energy subsidies around the world (IMF 2013). LAC “pre-tax” subsidies total $36 billion, about 0.5 percent of regional GDP and 7.5 percent of all global “pre-tax” energy subsidies. Across countries, subsidies for petroleum products are the largest in Ecuador and Venezuela, both major oil exporters (Table 3.7-5). No LAC countries subsidize coal and very few subsidize natural gas or electricity. Two LAC countries – Brazil and Chile – have undertaken successful reforms of energy subsidies since 1990 (IMF 2013). 3.7.3 Issues The adverse effect of subsidies on biodiversity and ecosystems are straightforward. As noted above, by boosting profits, they encourage targeted economic activities, which in turn harm biodiversity and ecosystems. For example, extending transportation networks can spur deforestation by increasing the profitability of agriculture and development near roads. Agricultural subsidies boost farm profits, which in some cases create incentives for farms, particularly large commercial ones, to clear forestland and expand their operations (Angelsen 1999; Binswanger 1991; Graham et al. 1986). Subsidies to agrochemicals may contribute to nonpoint source water pollution that harms freshwater and marine ecosystems (Myers and Kent 2001). Subsidies to high-yielding varieties can promote monocropping and loss of genetic diversity (Dyer and Belausteguigoitia 1996). Energy subsidies spur industrial, agricultural, and extractive activities, thus affecting biodiversity and ecosystems through urban expansion, pollution, and agricultural extensification. Not surprisingly, the main impediments to implementing subsidy reform are often political. Subsidies often have valid aims, including spurring economic development, mitigating poverty, and addressing concerns about income distribution. But political economy factors also shape subsidy policies, and beneficiaries usually fiercely defend them, which makes reform challenging. That said, when subsidies are driven mainly by pressures from special interest groups, cutting them is a win–win outcome: it benefits both the economy and the environment (Sterner 2003; Myers and Kent 2001; van Beers and de Moore 2001; IMF 2013). International competition in agriculture can create another significant obstacle to subsidy reform. Because many countries heavily subsidize agricultural products traded in international markets, unilateral subsidy reform can jeopardize exporters’ competitiveness



Table 3.7-5 Pre-tax energy subsidies as a percentage of 2011 GDP Country

Petroleum products


Natural gas


Antigua and Barbuda Argentina Bahamas Barbados Belize Bolivia Brazil Chile Colombia Costa Rica Dominica Dominican Republic Ecuador El Salvador Grenada Guatemala Guyana Haiti Honduras Jamaica Mexico Nicaragua Panama Paraguay Peru St. Kitts and Nevis St. Lucia St. Vincent and the Grenadines Suriname Trinidad and Tobago Uruguay Venezuela

0.49 0 0 0.04 0 2.4 0 0 0 0 0 0 6.31 0 0 0 0 n.a. 0.02 0 0 0 0.02 0 0 0.2 0.19 0 0 2.75 0 5.58

n.a. 1.03 n.a. n.a. n.a. n.a. n.a. 0 0 n.a. n.a. n.a. 0.18 0 n.a. n.a. n.a. n.a. n.a. n.a. 0 n.a. n.a. n.a. 0 n.a. n.a. n.a. n.a. n.a. n.a. 1.02

n.a. 0.77 n.a. n.a. n.a. n.a. n.a. 0 0 n.a. n.a. n.a. 0 0 n.a. n.a. n.a. n.a. n.a. n.a. 0 n.a. n.a. n.a. 0 n.a. n.a. n.a. n.a. n.a. n.a. 0.59

n.a. 0 n.a. n.a. n.a. n.a. n.a. 0 0 n.a. n.a. n.a. 0 0 n.a. n.a. n.a. n.a. n.a. n.a. 0 n.a. n.a. n.a. 0 n.a. n.a. n.a. n.a. n.a. n.a. n.a.

Source: IMF 2013.

(Peters 2006; van Beers and van den Bergh 2001). With regards to fisheries subsidies, negotiations have taken place at WTO meetings to ban perverse subsidies that enhance fishing capacity. Among the main issues to be resolved are: the differential treatment of developing and developed nations, transparency, enforcement, and scope (UNEP 2008). 3.7.4 Evidence Causal links between subsidies and resource degradation are generally straightforward, and evidence to support these links is available. For example, studies have linked fishing subsidies to overexploitation (Porter 2004; Sumaila et al. 2013). A considerable econometric literature shows that transportation networks, which are almost always state financed, exacerbate deforestation (Chomitz 2007). Finally, some evidence directly links LAC agricultural subsidies to deforestation. For example, in a study using data from nine countries (Costa Rica, the Dominican Republic, Ecuador, Honduras, Panama, Paraguay, Peru, Uruguay, and Venezuela), Bulte et al. (2007) find that subsidies led to both extensification



and (surprisingly) lower productivity, apparently because farmers had incentives to adopt inefficient, extensive modes of production to garner subsidies (see also Repetto 1988; Binswanger 1991). Similarly, Peres and Schneider (2012) argue that subsidized farmer relocation programs have caused significant deforestation in Brazil. They note that 70 percent of Brazilian Amazon land assigned for resettlement had been deforested by 2007 and that resettled lands accounted for 21 percent of all Amazonian deforestation in 2008.

3.8 Payments for environmental services 3.8.1 Description Payments for environmental services (PES), also known as payment for ecological services, are cash transfers from users of environmental services to providers of these services, conditional on continued provision. The term PES has been applied somewhat indiscriminately to a wide range of policies ranging from national park entrance fees to eco-certification (Engel et al. 2008). In a widely accepted definition, a true PES program entails (1) a voluntary transaction where (2) a well-defined environmental service (or a land use likely to secure that service) (3) is being “bought” by a (minimum one) service buyer (4) from a (minimum one) service provider (5) if and only if the service provider secures service provision (conditionality) (Wunder 2005). PES programs target a variety of ecosystem services, including carbon sequestration, watershed protection, landscape beauty, and protection of biodiversity habitat. 3.8.2 Status and trends Overall level of implementation LAC leads the developing world in PES implementation, possibly because land tenure is more secure than in Africa and Asia and because in most parts of the region, commercializing ecosystem services is culturally and politically acceptable (Southgate and Wunder 2009). In their compilation of PES (and PES-like) schemes worldwide, Landell-Mills and Porras (2002) found almost a quarter were in LAC (Table 3.8-1). Southgate and Wunder (2009) list 90 watershed PES schemes in LAC (Tables A3.8-S1 and A3.8-S2). Although PES is relatively common in LAC, four caveats are important. First, most PES schemes in LAC, as in other regions, are only PES-like and do not meet the five-point definition given above (Southgate and Wunder 2009). Second, most programs that have been planned or initiated, including those listed in Table 3.8-S1, have never actually gotten off the ground (Southgate and Wunder 2009). Third, many programs are relatively new and have Table 3.8-1 PES schemes, by region Region


Interregional LAC North America Europe Asia, Pacific Africa

28 24 17 14 10 7

Source: Landell-Mills and Porras 2002.



yet to be rigorously evaluated. Finally, most programs are local. Only a few LAC countries have significant national programs, notably Brazil, Costa Rica, Ecuador, and Mexico. Of these, the programs in Costa Rica and Mexico are the best established and best understood. Level of implementation by subregion Southgate and Wunder (2009) describe implementation of watershed PES programs – those focusing on programs supplying watershed services – by LAC subregion. They note that such programs, which are the most common type in LAC, have considerable promise in the Andes subregion because people live in mountainous areas with increasing water scarcity combined with considerable upland forest loss. Therefore, PES has the potential to stem scarcity by paying upland land managers to preserve forest cover, thereby providing hydrological services. However, actual implementation is uneven across the large Andean countries, with Ecuador and Colombia leading the charge and Bolivia, Peru, and Venezuela bringing up the rear. Ecuador has three of the oldest and best-known programs, namely PROFAFOR, Pimampiro, and Socio Bosque, a national program launched in 2008. By contrast, watershed PES potential in the Amazon and Brazil is more limited because water is more abundant, land is flatter, and the most heavily forested regions are lightly populated. That said, Brazil has offered Proambiente, a national program, since 2006, and a number of state-level programs including Bolsa Floresta, which was established in 2007. Finally, PES is relatively common in Mesoamerica. As discussed below, two of LAC’s oldest and largest PES programs, both national, are in Costa Rica and Mexico. 3.8.3 Issues The case for payments for ecosystem services Pure PES programs look very much like the theoretical solution to negative environmental externalities described by Coase (1960): that is, voluntary payments are made by victims to perpetrators conditional on the perpetrators’ mitigating or preventing the externality. The idea of this broad approach to conservation was partly spurred by the proliferation of integrated conservation and development programs of the 1980s and 1990s, which attempted to simultaneously encourage development and conservation by, for example, promoting commercialization of nontimber forest products (Pattanayak et al. 2010). Beginning in the late 1990s, a number of influential articles appeared asserting that direct payments for conservation would be simpler and more cost-effective and have the added benefits of tapping new sources of conservation finance (provided by victims of externalities) and alleviating poverty by channeling funds to small farmers and other suppliers of ecosystem services (Pattanayak et al. 2010). Barriers to effectiveness Observers have described at least ten reasons PES may fail to achieve the ends envisioned by early advocates (Pattanayak et al. 2010; Engle et al. 2008; Southgate and Wunder 2009; Alpízar et al. 2007; Wunder 2007; Wunder 2013): •

Lack of additionality due to self-selection. PES programs will not result in conservation beyond what would have occurred absent payments when participating ecosystem service suppliers disproportionately consist of land managers not planning to engage in


• •

Policies prohibited activities (clearing forest, logging, etc.) or those already planning to undertake required activities (sustainable agricultural practices, vegetative barriers, etc.). Such suppliers clearly have relatively strong incentives to volunteer into PES programs because their costs of doing so are relatively low. Leakage. PES, like many other conservation programs, can result in leakage – that is, the shifting of undesirable activities to adjacent areas. For example, a landowner who agrees not to harvest trees in one area may simply log nearby areas. Poor targeting. PES programs may do a poor job of targeting areas where conservation benefits like hydrological services are relatively high but costs, including opportunity costs of conservation, are low. In large-scale national programs where ecosystem service users are dispersed, targeting is a particular problem because political economy factors, such as poverty alleviation and a desire to distribute payments throughout the country, may have a strong influence on targeting. Poor monitoring and enforcement. Monitoring and enforcement of conditionality may be weak. For example, in many programs, the sanction for noncompliance is loss of future payments, not forfeit of past payments. Tenuous link to ecosystem services. Often the link between proscribed activities (e.g., land-use change) and the promised ecosystem services (e.g., reduced sedimentation and enhanced water yield) are tenuous. In the long term, the lack of a strong link reduces service buyers’ willingness to pay. Research has demonstrated that such links are heavily influenced by place-based factors that are difficult to assess rapidly and cheaply. In addition, the most important human activities contributing to degradation of ecosystem services may not be caused by land managers receiving payments, but by transients engaging in activities such as charcoal production and gravel mining. Market and government failures. Market or government failures may result in payments that are too small to alter the behavior of ecosystem service providers. One reason is that payments may only cover one among many bundled services. For example, payments might focus on watershed services and not carbon sequestration or preservation of biodiversity habitat. A second reason is that, in some cases, only a subset of ecosystem service users participate in the program. For example, in a watershed PES program, upstream suppliers may receive payments from only one among many communities of downstream users. A third reason is that government often subsidizes water, which reduces funds available for PES. Strategic behavior. In some cases, ecosystem service suppliers are few in number and therefore oligopolists may engage in strategic behavior to drive up the price of their product. Fixed costs. PES programs often entail significant fixed costs for activities such as design and administration. These costs may make small-scale programs impractical absent significant subsidies. Property rights. PES programs are difficult to administer when suppliers of ecosystem services lack well-defined individual or communal property rights. In such cases, identifying, enrolling, and contracting with suppliers is problematic. In addition, the distribution of property rights within a society – specifically who owns the land providing services and who owns the land receiving services – determine who receives the payments and therefore can create political and economic dynamics that undermine feasibility and implementation. Permanence. PES programs by definition entail continuous payments to suppliers and therefore will be neither permanent nor practical unless a long-term source of financing is identified.



Layered on top of all those challenges to effectiveness and efficiency are concerns about the distributional consequences of PES, particularly worries about whether poor households lacking well-defined property rights are able to participate, and whether and how nonparticipating households are indirectly affected by PES. 3.8.4 Evidence Although more than three dozen studies of PES programs in LAC have appeared (MartinOrtega et al. 2012), rigorous evidence in favor of their effectiveness is limited. Pattanayak et al. (2010) review this evidence for LAC and other regions and emphasize that, partly because PES is a relatively new approach, the literature is thin. That said, the authors identify eight rigorous credible evaluations of programs in Colombia, Mexico, China, and Costa Rica. They conclude that these studies demonstrate that “there are small effects of PES.” A 2009 study reviewing much of the same evidence reaches an even more negative conclusion, asserting that “the few studies that exist show little or no environmental impact” (GEF 2009). The two best known and most studied LAC PES programs – that feature most prominently in reviews just noted – are Costa Rica’s Payments for Environmental Services (Pagos por Servicios Ambientales, PSA) and Mexico’s Payment for Environmental Hydrological Services (Pago de Servicios Ambientales Hidrológicos, PSAH). There is an emerging consensus in the literature that, at least in their early incarnations, neither program was effective at targeting forested areas that both provide important environmental services and face a significant risk of deforestation, and that neither program had large additional effects on avoided deforestation (Blackman and Woodward 2010). The main component of Costa Rica’s PSA program aims to provide hydrological benefits, biodiversity, and other environmental services by paying managers of forested lands to retain forest cover. Yet as of 2005, only 35 percent of the land participating in the PSA program was in a watershed with downstream users of hydrological services, and depending on the definition of biodiversity priority areas, 30 to 65 percent of PSA land was in biodiversity priority areas (Pagiola 2008). Targeting of payments to land that provides important ecosystem services has improved over time (Pagiola 2008; Barton et al. 2009). However, the challenge of targeting for additionality remains. Virtually all rigorous statistical analyses based on forest cover data derived from satellite images find that the PSA program has done little to avoid deforestation, largely because land at high risk of deforestation has not been volunteered into the program. Instead, the lion’s share of land enrolled in the program has been ill-suited for agriculture, pasture, or other cleared land uses and very probably would have remained forested absent the program (Robalino and Pfaff 2013; Robalino et al. 2008; Arriagada et al. 2012; Sills et al. in press).3 Mexico’s PSAH program shares many of PSA’s design elements, including a focus on forest conservation and voluntary enrollment. However, as its name suggests, the program aims specifically at ensuring the provision of hydrological benefits. Also, its administrators have made an effort to target high-benefit areas. Evaluations of the early years of the program suggest that initial targeting efforts were disappointing. In 2006, 51 percent of land enrolled in the program was in watersheds classified as “not overexploited,” and 68 percent was deemed to have low or very low deforestation risk (Muñoz-Piña et al. 2008; AlixGarcia et al. 2005). As a result, the program’s effect on conservation has been modest (Alix-Garcia et al. 2012; Honey-Roses et al. 2011).



3.9 Eco-certification 3.9.1 Description Eco-certification programs, also known as eco-labeling and sustainability certification programs, accredit goods and services that have met defined process standards meant to protect the environment and social welfare (Blackman and Rivera 2011). They are administered by governments, NGO, trade associations, multilaterals, and even individual companies, and they focus on a wide variety of commodities including bananas, cocoa, coffee, and timber, final goods like furniture and paper, and services like tourism. Process standards can cover an array of factors. For example, Forest Stewardship Council (FSC) certification requires adherence to ten broad principles (as well as more specific criteria associated with each principle): (1) complying with applicable national laws; (2) defining and respecting longterm tenure and use rights; (3) respecting indigenous peoples’ rights; (4) enhancing the longterm social and economic well-being of local people; (5) encouraging the efficient use of the forests’ multiple products and services; (6) minimizing environmental impacts to conserve biodiversity, water resources, and ecosystem services; (7) developing a written forest management plan; (8) monitoring forests, management activities, and their social and environmental impacts; (9) maintaining high conservation value forests; and (10) planning plantations in accordance with the above principles (FSC 1996). 3.9.2 Status and trends Eco-certification programs are increasingly popular worldwide. More than 430 labels have been established in 197 countries and 25 industrial sectors (Ecolabel Index 2013). Ecocertification of agricultural commodities is particularly widespread. Today, 10 percent of the timber, 7 percent of the coffee, and 12 percent of the wild fish products traded in international markets are eco-certified (Eilperin 2010). Data on the extent of eco-certification by geographic region is not publicly available. A recent report that attempted to compile these data indicates that LAC is global leader in eco-certification. It produces 97 percent of the world’s certified bananas, 75 percent of its certified coffee, and 48 percent of its certified cocoa (Potts et al. 2010).

3.9.3 Issues According to advocates, eco-certification causes producers to adopt greener practices (Giovannucci and Ponte 2005; Rice and Ward 1996). In theory, the causal chain begins when eco-certification enables consumers who prefer green goods to identify and purchase them. Elevated demand for such goods generates price premiums and better market access for green producers, which in turn creates financial incentives for them to hew to sustainable production practices. More important, it encourages relatively dirty producers to adopt sustainable practices so that they can get eco-certified. If that logic holds, eco-certification could be an important tool for addressing environmental problems that threaten ecosystems and biodiversity in Latin America. Growing and processing bananas, timber, coffee, cocoa, and other agricultural products in Latin America causes deforestation, soil erosion, and agrochemical pollution. These problems are quite difficult to tackle using conventional top-down government regulation because producers are often small, numerous, and geographically dispersed, while regulatory institutions are



undermanned and underfunded (Wehrmeyer and Mulugetta 1999). Eco-certification schemes have the potential to sidestep these constraints by creating a private sector system of economic incentives, monitoring, and enforcement (Cashore et al. 2004; Dingwerth 2008). But not every eco-certification program will have these benefits (Blackman and Rivera 2011). To do that, a program must set and enforce standards stringent enough to ensure that dirty producers are excluded. In addition, it must generate price premiums high enough to offset the costs producers pay to meet certification standards and to deal with the red tape involved in getting certified. Even if these two conditions are met, program benefits still can be undermined by so-called selection effects. Producers already meeting certification standards have strong incentives to select into eco-certification: they do not need to invest in adopting sustainable practices to pass muster and can obtain price premiums and other benefits. But programs that mainly attract already-green producers will not change producer behavior and therefore will have limited environmental benefits. Because there are good reasons to expect eco-certification to have significant environmental benefits in developing countries, and good reasons to expect it does not, empirical research is needed to figure out whether and in what cases benefits actually are realized. 3.9.4 Evidence Rigorously assessing the environmental effects of eco-certification can be challenging because as just noted, already-green producers have incentives to disproportionately select into certification programs and because factors other than eco-certification, including technological progress, can improve producers’ environmental performance. A failure to control for these factors can generate gross overestimates of the benefits of eco-certification. Recent literature reviews find that few studies have tried to measure the producer-level environmental effects of eco-certification, and even fewer have overcome, or even tried to overcome, the methodological challenges. Blackman and Rivera (2011) present a comprehensive review of the literature on the producer-level environmental and socioeconomic effects of eco-certification, with an emphasis on methodological rigor. The authors found just 46 peer-reviewed studies, which collectively have very limited scope. The bulk of the studies examined just two sectors (coffee and forest products) and two certifications (Fair Trade and the FSC). Moreover, of these 46 studies, only 11 used methods likely to produce unbiased results. And of these 11 studies, just two focused on the environmental (versus socioeconomic) effects of certification. Neither found a significant benefit. Another recent comprehensive review (ITC 2011) generated similar qualitative results. Happily, since these reviews appeared, the pace of evaluations of eco-certification impacts seems to have increased somewhat (e.g., Subervie and Vagneron 2013; Blackman and Naranjo 2012; Blackman et al. 2012; Ruben and Fort 2012; Gutierrez et al. 2012). Nevertheless, we still do not yet have nearly enough evidence to determine whether and under what conditions eco-certification is effective in protecting biodiversity.

3.10 Ecotourism 3.10.1 Description Ecotourism is “responsible travel to natural areas that conserves the environment and improves the welfare of local people” (TIES 2012). According to the Center of Ecotourism and Sustainable Development, ecotourism (1) involves natural areas; (2) minimizes environmental



impacts; (3) helps educate and facilitates interaction between the traveler and the host community; (4) builds environmental awareness; (5) provides direct financial benefits for conservation; (6) provides financial benefits and empowerment for local people; and (7) respects local culture, human rights, and democratic movements (Honey and Krantz 2007). 3.10.2 Status and trends Tourism – not necessarily ecotourism – is one of the largest and fastest-growing sectors in LAC, often in coastal areas. It accounts for 6.6 percent of GDP in the region, and in some subregions, especially the Caribbean, it accounts for far more. From 1995 to 2007 tourist arrivals to Latin America grew 49 percent, and tourist receipts grew from $2.3 billion to $3.7 billion (Fayissa et al. 2009). LAC tourism is closely linked to nature. Nearly all Caribbean tourism and hospitality companies have indicated in surveys that their business relies on the surrounding natural environment (Slinger 2002). More than half of international tourists to Peru, Argentina, and Costa Rica report that nature and wildlife were their main reason for visiting. In Costa Rica, more than half of tourists visit at least one protected area. Ecolodges, an indicator of ecotourism activity, are plentiful in LAC’s most biodiversity-rich countries, including Costa Rica (590), Peru (356), Ecuador (345), and Mexico (304) (Bovarnick et al. 2010). Originating in the 1970s and 1980s, ecotourism as a concept has grown in popularity and prominence over the past three decades. By 2005, three years after the UN declared the International Year of Ecotourism, it was a $540 billion global market with an annual growth rate of 5 percent (Honey and Krantz 2007). In LAC, Costa Rica has been a leader in the development of ecotourism. Although multilateral and bilateral organizations scaled back their tourism programs in the 1970s, partly because tourism was perceived to be an unreliable economic development tool, since the 1990s they have increased their support for ecotourism projects (Honey and Krantz 2007). NGOs like World Wide Fund for Nature (WWF) also have promoted ecotourism. 3.10.3 Issues and evidence The term ecotourism often is used indiscriminately and applied to tourist activities that are not particularly environmentally friendly. Ecotourism certification has been advocated as a solution to this problem. However, evidence that this approach is successful is lacking, partly because eco-certification standards and enforcement vary widely (Blackman and Rivera 2011; Buckley 2002; Font and Buckley 2001; Buckley 2011). A broader and more important issue is that even ecotourism activities meeting reasonable standards can have both positive and negative effects on conservation. As for positive effects, tourism can promote the creation of new public and private protected areas and better management of existing PAs. In principle, tourism operations that depend on nature in and around PAs have an incentive to promote and even finance them. In addition, flourishing ecotourism arguably creates economic incentives for regulatory authorities to conserve biodiversity, generates the funding they need to do that, and educates the public about the benefits of biodiversity and ecosystem services (Steven et al. 2013; Buckley 2010, 2011). But ecotourism can also destroy and degrade habitat, introduce invasive species, and cause pollution. It increases human presence in natural areas, which in turn can lead to new economic activities and the building of new roads, sewers, electricity, and other infrastructure, particularly on the edges of PAs (Steven et al. 2011; Pickering and Hill 2007; Buckley 2004, 2011). Evidence from case studies of terrestrial protected



areas in Costa Rica and Belize suggest that there is significant variation in the range of effects that human visitors have on both flora and fauna in protected areas (Farrell and Marion 2001). Marine protected areas can also be sensitive to human visitors, with consequences for corals as well as aquatic vegetation (Milazzo et al. 2002). As with other conservation policies, the net effect of ecotourism on local biodiversity depends critically on place-based factors (Wunder 2000).

3.11 Bioprospecting 3.11.1 Description Bioprospecting refers to the systematic collection of biological samples, taken from in-situ plants and animals, for their genetic and biochemical information as sources of new or improved products and processes. It is most commonly associated with pharmaceutical development. For example, Epibatidine, a powerful pain medication, was derived from material found in Ecuadorian tree frogs, and Tubocurarine, a muscle relaxant, was derived from a South American vine (Quezada 2007). Bioprospecting also contributes to a wide variety of other productive uses. For example, natural enzymes are used in pulp and paper production, oil drilling, and textile manufacturing. Bioprospecting is typically carried out by consortiums of multinational corporations, universities, and research centers. 3.11.2 Status and trends A recent study of bioprospecting activities in LAC reached the following conclusions (Quezada 2007). First, relatively few firms dedicated exclusively to bioprospecting are operating in LAC, probably because heavy capital requirements constrain entry. Second, LAC research institutions have been successful at establishing alliances with bioprospecting companies and, as a result, have developed robust in-house capacity and infrastructure. However, technology and capital requirements in bioprospecting are advancing quickly, making it difficult for these research institutions to keep up. And finally, access and benefit sharing (ABS) rules vary widely across LAC countries. Such rules dictate how bioprospectors can access biological materials and how the proceeds are shared with governments and local communities. The first formal ABS agreement in LAC was negotiated between Merck, the pharmaceutical multinational, and INbio, a Costa Rican research institute. By 2004, nine LAC countries had put in place ABS legislation (Argentina, Brazil, Costa Rica, El Salvador, Ecuador, Mexico, Nicaragua, Panama, and Peru), and ABS principles were set forth in a directive of the Andean Community (comprising Bolivia, Colombia, Ecuador, Peru, and Venezuela). Brazil is now widely seen as being at the forefront of LAC ABS rules. Various initiatives, including the International Cooperative Biodiversity Group, are helping to promote ABS rules in the region. 3.11.3 Issues Two intertwined issues related to bioprospecting have stirred considerable controversy. The first is the expected per hectare return. According to some, returns can provide strong incentives for conservation if benefits are shared with local land managers (Reid et al. 1993; Rubin and Fish 1994). In principle, such benefits could help remedy underinvestment in conservation that ultimately stems from the fact that biodiversity is an open-access resource. We return to this issue in the next subsection.



A related issue, of course, is how benefits are shared. Bioprospecting has created disputes over compensation between the bioprospectors (usually large companies), the owners of the land being prospected (mostly private or government entities), and the inhabitants of this land (often poor forest dwellers). Such conflicts have prompted the development of the ABS agreements discussed above. Unfortunately, in LAC, these agreements have not been particularly successful at resolving benefit-sharing issues because of conflicts with other rules and regulations, institutional weakness, and political economy considerations. According to Quezada (2007: 31), Despite the ongoing efforts of host nations to facilitate access and sharing of benefits and biodiversity conservation, experiences to date have demonstrated that many of the prevailing regulatory regimes have instead constituted roadblocks to arrangements and deals that would be of value to all parties. Stromberg et al. (2013) reach similar conclusions. 3.11.4 Evidence 4 Evidence clearly demonstrates that in-situ plants and animals are an important source of valuable new products and processes. For example, in the context of health care, Newman et al. (2003) analyze new drugs approved by regulatory agencies from 1981 to 2002 and find that most had origins in nature. According to this study, 57–62 percent of all novel, active substances for treating disease are initially derived from natural biological sources. These include approximately 62 percent of cancer medications and about 65 percent of drugs for treating hypertension. There is some variation across disease categories, but fully synthetic drugs are generally a minority of new drugs. Nevertheless, the economic value of bioprospecting options alone is unlikely to provide sufficient incentives for the preservation of biodiversity. Simpson et al. (1996) demonstrate that when researchers must analyze an abundance of biological material to locate a single species that yields valuable material, the average incremental value of any one species in an ecosystem is not very high, particularly if several species are close substitutes. The economic returns from bioprospecting per hectare are likely quite modest, even in biodiversity hotspots that are rich in endemic species – from $0.20 per hectare in the California Floristic Province to $20.63 in western Ecuador – and even under assumptions that likely overstate their option benefits. For this reason, the potential for new product development alone usually does not provide enough benefit to compel private parties to invest in biodiversity conservation (Simpson et al. 1996; Barrett et al. 2013; Costello and Ward 2006). This conclusion is supported by the fact that bioprospecting has thus far focused principally on pharmaceutical development, which suggests that for other uses, the benefits of bioprospecting have rarely exceeded the opportunity costs.

3.12 Mitigation offsets and banking 3.12.1 Description Mitigation offsets are ecosystem improvements that land managers undertake in one location in order to compensate for ecosystem damages elsewhere (US EPA 1995; Ruhl et al. 2005). For example, landowners that drain natural wetlands to build residential developments



may construct wetlands in an adjacent area to compensate for the ecological damage, and mining companies that clear forest for their operations may fund afforestation projects in nearby areas. Offsets generally are mandated by regulations but can also be voluntary. In regulatory systems, several variations are available for implementation, including habitat banking and tradable development rights. These options are often one among a hierarchy of three broad regulatory strategies aimed at limiting environmental losses from development projects: avoidance, mitigation, and compensation. Offsets typically entail restoring more acres than are disturbed in order to account for the fact that not all hectares of habitat provide the same level of ecosystem services and to ensure additionality. Offsets can be operationalized in two ways. Development projects can be required to undertake them themselves. Alternatively, they can purchase offsets from mitigation private or public banks serving multiple developers. Banks undertake mitigation projects and provide offsets, also referred to as credits, to developers (US EPA 1995, 2009; US DOI 2003). 3.12.2 Status and trends Although mitigation offset and banking are subject to growing interest in LAC, they have yet to be formally implemented in the region (Bovarnick et al. 2010; Madsen et al. 2011). Mitigation offsets Within LAC, regulations to reduce the environmental impacts of infrastructure and other public- and private-sector economic development projects generally occur in the context of environmental licensing. Hence, mitigation offsets could be linked to licensing. For example, Colombia is in the process of establishing a mitigation hierarchy process within environmental licensing that includes mitigation offsets (TNC 2011). In Colombia the legal foundation for offsets has been in place since 1993. By law, environmental licenses require prevention, mitigation, correction, compensation and management of a project’s environmental effects. A resolution passed in 2010 specified that compensation for biodiversity loss must be undertaken according to rules being developed cooperatively by the Ministry of Environment, The Nature Conservancy, World Wildlife Foundation, and Conservation International. These rules specify areas where projects will require mitigation, the location where offsets should occur, and the specific way in which the required quantity of offsets will be calculated. Other countries, including Chile and Brazil, have expressed interest in this approach (A. Ramos, personal communication). Mitigation banking A recent report by UNDP (Bovarnick et al. 2010) evaluates the potential and feasibility of habitat banking in LAC, but mitigation banking has yet to be implemented. A brief discussion of its history in industrialized countries sheds light on its potential for LAC. Mitigation banking originated in the United States, where two main types of banks exist: wetland banks, which serve to achieve the goal of “no overall net loss” of wetlands; and conservation banks, which are designed to mitigate the depletion of nonwetland habitats and the protection of rare or endangered species. Wetland mitigation banking and conservation banking more generally have their origins in the compensatory mitigation of wetlands mandated by the 1972 Clean Water Act in the United States. The use of mitigation banks did not start until 1983, when the US Fish and Wildlife Service issued guidance on wetland mitigation banking.



The first wetland mitigation banks were managed by public agencies or large corporations (Bean et al. 2008). The first commercial mitigation banks were established in the early 1990s, and proliferated rapidly after the Environmental Protection Agency and the Corps of Engineers released banking guidance in 1995. Today, roughly 900 banks exist, accounting for more than half a million acres of wetlands (Madsen et al. 2011). Conservation banking in the United States began at the state level. In 1995 California became the first state to allow for it (ELI 2009). Formal federal guidance for the use of conservation credits to meet the mitigation requirements under the Endangered Species Act were issued in 2003. Currently, about 75,000 acres is permanently protected by a total of 109 conservation banks. On average over the past ten years, about 4,400 acres has been added annually to the conservation banks, the vast majority in California (Madsen et al. 2011). Mitigation banking is also used for compensating ecological loss due to habitat development outside the United States, but the scope of mitigation activities and banking is more limited, except perhaps in Australia, where several biodiversity offset and mitigation banking programs are in place, including the Bush Broker program in Victoria, the Bio Banking program in New South Wales, and several other active programs elsewhere in the country (Madsen et al. 2011). Canada also has mitigation banking, primarily at the provincial level (Madsen et al. 2011). Finally, mitigation banking and, more broadly, biodiversity offsets, are gaining increased attention in Europe but remain at planning and pilot project stages (Madsen et al. 2011). For example, the European Union’s strategy for reversing biodiversity loss identifies biodiversity offsets as one of the potential innovative mechanisms to achieve biodiversity conservation goals and to increase private-sector involvement and funding. Tradable development rights Tradable development rights (TDRs) facilitate ecological mitigation by creating a market that allows development rights to be transferred from one location to another. The regulatory agency sets a minimum amount of undeveloped (protected) area, which is equivalent to a cap on the total area that can be developed. Developers are allowed to bypass the cap by purchasing development rights from other locations. Areas where rights are purchased then enjoy increased protection. Many TDR programs have been established in the United States over the past 30 years, but only a few have been broadly successful in achieving local land-use goals (McConnell and Walls 2009). One of the main difficulties with TDRs is the complexity of integrating markets with local land-use planning. In addition, TDR markets heavily depend on local economic, housing, and land market conditions and their variations. Regardless, there is great potential under some conditions. Local governments may find TDRs appealing because they help protect land without relying on tax revenue or necessarily limiting overall growth in a community. Landowners and developers tend to view TDRs positively because they rely on markets and are voluntary. Economic grounds also support TDRs; they can help achieve conservation goals at a reduced overall cost to the local economy (Messer 2007; McConnell and Walls 2009). Experience with TDRs in LAC is limited, but they have been an important policy instrument in Brazil (Karsenty 2007). The Brazilian Forest Code requires landowners to protect a certain minimum percentage of land, and the forest reserve system allows trading between landowners to meet their obligations to protect forest cover. The required protection percentage varies by location and type of land. For example, landowners within the legally defined forested Amazon are required to protect 80 percent of land, while those in other parts must



protect only 20 percent (EFTEC-IEEP et al. 2010). TDRs have many conceptual similarities to habitat banking, and a clearinghouse is required for managing both systems. 3.12.3 Issues Mitigation offsets have the potential to reduce the conflict between economic development and biodiversity conservation. This is particularly true of mitigation banking, which creates a market for meeting regulatory requirements. This, in turn, can provide opportunities to greatly reduce the cost of meeting regulatory requirements. Furthermore, unlike regulatory approaches that place the burden on developers and landowners to meet the environmental goals on their property, mitigation banking lets them pass this responsibility onto third parties, which may have greater interest and experience in successful conservation. On the other hand, several issues remain challenging in the design and implementation of mitigation banks, such as how to identify the most environmentally preferable offsets within a landscape context and determine appropriate mitigation replacement ratios (Ruhl et al. 2005; McKenney and Kiesecker 2010). Mitigation banking in LAC has several economic and institutional prerequisites (Bovarnick et al. 2010): • •

• • •

policies or regulations aimed at reducing biodiversity loss due to habitat development or resource extraction; integration of banking into land-use permitting and/or environmental impact assessments that require compensation of ecological losses due to development or resource extraction; demand for, and supply of, mitigation credits created by robust regulatory drivers of mitigation banking; government and third-party management of mitigation banking programs; and scientific and market support services to provide implementation, monitoring, legal, and registry services.

Using those criteria, Bovarnick et al. (2010) assesses the feasibility of mitigation banking in eight LAC countries: Argentina, Brazil, Colombia, Costa Rica, Chile, Mexico, Panama, and Peru. They find that banking is feasible in each case. Each country has national-level initiatives and strategies in place to limit habitat loss, each has permitting and environmental impact assessments that could be adapted to facilitate banking, and each possesses the scientific and technical capacity needed to provide support services to mitigation banking programs. Notwithstanding that optimistic assessment, Bovarnick et al. (2010) also see significant barriers to the use of offsets and mitigation banks in LAC. Although the necessary national environmental regulatory frameworks are generally in place, enforcement and implementation are typically limited and institutions are often weak (see Section 3.6). In addition, experience in the United States suggests that the startup costs of mitigation banks are high, and it is often necessary to offer initial credits before a bank is established (Robertson 2006). Given such realities, Bovarnick et al. (2010) conclude that only four of the eight countries included in the report are ready to move ahead with mitigation banking in the short term: Brazil, Costa Rica, Chile, and Mexico. Generally, the report recommends that mitigation banking be carefully developed, with stakeholder consultations, pilot projects, capacity building, and regulatory reforms as the next steps throughout the region before moving ahead with larger-scale efforts to implement mitigation banking (Bovarnick et al. 2010).




3.13 National environmental accounting 3.13.1 Description National environmental accounting, also known as green accounting, entails systematically collecting and disseminating data on environmental quality and natural resources and, ideally, incorporating them into national income measures, such as GDP (Asheim 2000). Specifically, it aims to incorporate information on the stocks of natural resources, the generation of pollutants and solid wastes, and environmental protection and restoration expenditures. Environmental accounting generates measures of “sustainable income,” typically defined as the maximum consumption (income) possible in the present period given the requirement that consumption in future periods must not decrease (following Hicks 1946). In principle, environmental accounting – and perhaps just as important, the data collection framework that underpins it – could help improve the effectiveness of biodiversity conservation policies by facilitating conservation planning and evaluation and by enhancing awareness of biodiversity issues in civil society. For example, the accurate, standardized panel data on fish stocks that are needed for environmental accounting also are needed to plan and evaluate marine biodiversity conservation policies. 3.13.2 Status and trends In LAC, national environmental accounting was first implemented in the late 1980s in Costa Rica (via a World Resources Institute pilot project) and in Mexico (via a World Bank– United Nations–STAT pilot project). Since then, ten other countries – all but one in Latin America – have implemented pilot projects (Table 3.13-1). Today, Peru, Colombia, Mexico, and Guatemala all have active, full-fledged programs (Edens 2012). However, few countries in LAC or any other developing regions have experimented with environmental accounts that include data on ecosystem services other than land use and land cover (Edens 2012).5 The programs in Colombia and Mexico often are singled out as success stories. Colombia has developed flow accounts for pollutants and materials as well as environmental indicators Table 3.13-1 LAC countries with environmental accounting programs, 2012 Country


Argentina Bolivia Brazil Chile Colombia Costa Rica Dominican Republic Guatemala Mexico Panama Peru Venezuela

9 9 9 9 9 9 9

Source: Edens 2012.

9 9

Regular program

9 9 9 9



for air and water quality and environmental protection expenditure accounts (INTOSAI 2010; World Bank 2006). Mexico’s environmental accounting program is one of the most complete in any developing country. Its green accounts cover, among other things, pollutants, natural resource stocks, and land use, and are used to calculate “green gross domestic product.” Green GDP, in turn, has been used as an economic indicator in national development plans for 2001–2006 and 2007–2012, which implies that the program is having real policy impacts (Edens 2012). A key to the program’s relative success has been Mexico’s framework environmental legislation, the 1988 General Law of Ecological Equilibrium and Environmental Protection (LGEEPA in Spanish), which requires calculation of green GDP. 3.13.3 Issues How to measure sustainable income has prompted considerable debate. Two distinct approaches have been advocated (Pearce and Warford 1993; World Bank 2006). The first involves measuring a country’s environmental quality and natural resource stock, placing monetary values on these measures, and using these values to adjust conventional measures of national income in some way (Repetto et al. 1989; Solarzano et al. 1991). The second approach – often called the physical accounts method – entails keeping parallel environmental and natural resource accounts denominated in physical units (e.g., volume or weight) rather than in monetary units. This approach has been explored most fully in Norway and France (Alfsen and Lorentson 1987; Theys 1989). Each approach has advantages and disadvantages. The first approach facilitates comparing environmental and economic data and creating a single indicator of sustainable income. The principal disadvantage is the wellknown difficulty of developing monetary values for nonpriced environmental goods and services. This can be particularly challenging in the case of biodiversity, which provides an array of valuable ecosystem services (Hamilton 2013). Views also differ as to how to adjust conventional aggregate income accounts to reflect the use of natural assets. One approach is to adjust net domestic product (NDP) to include the depreciation of natural assets as well as human-made capital (Bartelmus and van Tongeren 1993; Pearce and Warford 1993). However, some researchers consider this approach flawed because it ignores the permanent income that can be generated from the sale of natural assets; they advocate modifying GDP directly (El Serafy 1989). Controversy also surrounds the appropriate treatment of defensive and restorative expenditures. Some analysts argue that because these expenditures are costs rather than final products, they should be deducted from conventionally calculated GDP (Daly 1989; Pearce et al. 1990). Others argue that this approach is unlikely to yield useful results, given the difficulties of defining defensive expenditures (Bartelmus 1992). Finally, there has been some debate regarding the best method for structuring sustainable income accounts. Some researchers argue that environmental data should be integrated into a single measure (Hueting 1980; Peskin and Lutz 1990; Daly 1989). Others advocate keeping environmental data in a satellite account to preserve the statistical continuity of conventional income measures (Bartelmus 1992). As for factors affecting implementation of environmental accounting, a 2012 World Bank study identifies enabling conditions for three of the most successful programs in the developing world – those in Mexico, Colombia, and South Africa – and obstacles that hindered programs in other countries (Edens 2012). All three successful cases were middle-income countries that already had well-developed national income accounts. Programs in each country had a steering committee with engaged stakeholders and a clear implementation



plan, were staffed with capable personnel, and enjoyed support from prominent government figures – either politicians or heads of agencies. Major obstacles to successful environmental accounting programs in other countries included a lack of support, leadership, and funding and an absence of high expectations for the quality of data gathered. 3.13.4 Evidence To our knowledge, studies of how national green accounting affects environmental and natural resource management do not yet exist (U. Narain, World Bank WAVES program, personal communication).

3.14 Corporate social responsibility 3.14.1 Description Corporate social responsibility (CSR), also known as “beyond compliance behavior,” refers to actions not required by law that farms and firms take to improve environmental quality, workers’ health and safety, and community welfare (Lyon and Maxwell 2008a; Portney 2008). In this book, we focus on biodiversity CSR, which is CSR aimed at conserving biodiversity by, for example, reforesting cleared land, adopting sustainable agricultural practices, and restoring degraded wetlands. CSR is generally broken down into actions that firms take (1) unilaterally; (2) in the context of a voluntary program administered by a government, NGO, or multilateral organization; and (3) as part of a voluntary agreement negotiated with a governmental organization. In developing countries, the second type of CSR has received the most attention, and in the environmental and biodiversity realm, a few international voluntary programs dominate the landscape: IUCN’s Business and Biodiversity Program and the UN’s Global Compact program (a general program that devotes three of its ten guiding principles to environmental management). IUCN (2012) and WBCSD (2010, 2012) compile dozens of cases studies of biodiversity-related CSR, including some in LAC. 3.14.2 Status and trends Biodiversity CSR is increasingly widespread worldwide, including in LAC (Bhattacharya and Managi 2013; Kitzmueller and Shimshack 2012; WBCSD 2010, 2012; IUCN 2012). However, comprehensive, comparable data on CSR at the regional- or even country-level is scarce, in part because CSR is often unilateral and therefore disparate and uncoordinated. To our knowledge, Haslam (2004) is the only effort to characterize CSR in LAC, although several other papers compile case studies from LAC and list CSR participants (e.g., UNCTAD 2010; Correa et al. 2004). Overall levels of corporate social responsibility, by type of participant Haslam (2004) presents the results of a 2003 study of CSR (of all types, not just biodiversity related) in LAC based on a review of websites of the top 50 companies in each of 13 LAC countries and, for comparison’s sake, in two industrialized countries, Canada and the United States. He ranks participation in each of the 15 countries from high to low, broken down by



three types of participation: private sector, government, and public awareness. He finds a “huge gap” between the overall levels of CSR in industrialized countries and LAC countries (Table 3.14-1). Within LAC, countries fall into three CSR categories: three highly developed LA countries that are “catching up” with Canada and the United States (Argentina, Chile, and Mexico); four countries that are “walking” (the rest of the South American countries – Bolivia, Colombia, Peru, and Venezuela); and six countries that are “stalled” (the Central American and the Caribbean countries – Costa Rica, Nicaragua, Cuba, the Dominican Republic, Jamaica, and Trinidad and Tobago). As for the three types of participation in CSR in LAC, private-sector participation – that is, CSR activities undertaken by firms themselves – is weak compared with the two industrialized countries. In LAC, most CSR activities are undertaken by industry and trade associations representing individual firms, not by the firms themselves. Government participation in LAC is also relatively weak. However, public participation in LAC is high, at least relative to firm and government participation. International pressures A salient feature that distinguishes CSR in LAC and other regions of the global South from CSR in industrialized countries is the strong influence of foreign actors (Haslam 2004; UNRISD 2000; Barkin 1999). In industrialized countries, domestic actors, including firms themselves, governments, trade associations, NGOs, universities, and labor organizations, are the main drivers of, and participants in, CSR. In LAC, however, leading actors include headquarters of transnational corporations that pressure their overseas affiliates to adopt Table 3.14-1 Corporate social responsibility in selected countries in the Americas, by type of participation, 2003 Country Caribbean Cuba Dominican Republic Jamaica Trinidad and Tobago Central America Costa Rica Nicaragua South America Argentina Bolivia Chile Colombia Peru Venezuela North America Mexico Canada United States Source: Haslam 2004.

Private-sector participation


General public awareness

None None None None–low

None None None–low None

None–low None–low None None

None–low None

None None

None–low None–low

Low–medium None–low Medium None–low None None

None–low None Low None–low None None–low

Medium None–low Medium Low Low None–low

Medium Medium–high Medium

Medium High Low–medium

Low–medium Medium Low–medium



CSR, international NGOs, private foundations, foreign governments, and multilaterals. Several multilateral organizations are particularly important (Haslam 2004): •

• •

OECD. The Organisation for Economic Co-operation and Development (OECD) has developed voluntary CSR guidelines for transnational corporations and required member states, including Mexico and Chile in LAC, to create national points of contact, which are charged with promoting CSR guidelines. It also encourages aspiring OECD member states in LAC and elsewhere to establish national points of contact. However, the points of contact are typically understaffed and underresourced. IDB. The Multilateral Investment Fund of the IDB, along with the Organization of American States, has since 2002 sponsored annual conferences on CSR, now known as CSRAmericas. This effort grew out of a formal plan of action adopted at the 2001 Summit of the Americas meeting, which called for more CSR in the region. United Nations. The United Nations promotes CSR through its Global Compact program, a code of conduct that has been promoted by the UN Development Program. World Bank. The World Bank promotes CSR through, among other groups, the World Bank Institute.

In addition to multilaterals, several other types of foreign institutions have promoted CSR in LAC. These include international NGOs and parties to international agreements. Among the most prominent are the IUCN, which as noted above offers the Business and Biodiversity Program, the World Business Council for Sustainable Development, and the Convention on Biodiversity, which in 2006 adopted Decision VIII/17 aimed at furthering involvement of the private sector in the convention’s activities (Bled 2009; Young and Limjirakan 2011). 3.14.3 Issues The case for corporate social responsibility According to advocates, CSR has the potential to overcome well-known obstacles to the effectiveness of conventional, top-down, public sector-led approaches to biodiversity conservation in LAC, such as protected areas and land-use planning. These obstacles include gaps and inconsistencies in written regulations, weak environmental regulatory agencies, a preponderance of difficult-to-monitor small and informal firms and farms, and perhaps most important, a lack of political will to allocate scarce resources to environmental protection and to enforce environmental regulations. On its face, CSR appears able to sidestep these constraints because it does not depend directly on the public sector to issue mandates, monitor compliance, and sanction violations. Rather, it is driven by other pressures (Kitzmueller and Shimshack 2012; Portney 2008; Lyon and Maxwell 2008b; Koehler 2008). Specifically, consumers who prefer final and intermediate goods produced in an environmentally sustainable manner create market incentives for CSR. Creditors, including shareholders, may prefer lending to firms that undertake CSR. Employees who benefit from health and safety CSR or who prefer green goods themselves exert pressure through the labor market. NGOs exert pressures through media, markets, and political channels. And finally, at least some CSR is privately profitable, which of course creates economic incentives. Hence, some have argued that CSR can help promote conservation even in countries where regulatory institutions are weak.


71 Barriers to effectiveness Others argue that CSR is unlikely to become as widespread or deeply engrained in LAC as it is in industrialized countries because – notwithstanding pressures from transnational, multilateral, and other foreign entities, which as noted above are not insignificant – the domestic drivers of CSR are relatively weak in most LAC countries, for four reasons (Blackman 2010, 2008; UNRISD 2000). First, many of the nonregulatory factors that drive CSR are relatively weak. For example, niche markets for green products are smaller than in industrialized countries; capital markets, including stock markets, are thinner; and environmental NGOs and advocacy groups are relatively weak and scarce (Fry 1988; Wehrmeyer and Mulugetta 1999). To be fair, however, some evidence suggests that capital markets in developing countries are increasingly concerned with firms’ environmental performance (Gupta and Goldar 2005; Dasgupta et al. 2001, 2006). In addition, some suggest that niche markets for green goods are growing (Potts et al. 2010). Second, regulatory pressures, which are often frail in LAC, turn out to be a major driver of CSR. Considerable research suggests that firms participate in CSR when they believe that a failure to do so may trigger more stringent mandatory regulation (Lyon and Maxwell 2008b; Koehler 2008). Third, small-scale firms and farms are more prevalent in LAC than in industrialized countries (Peres and Stumpo 2000; Blackman 2006). These firms may be less susceptible to at least some regulatory and nonregulatory pressures for CSR, including those generated by capital markets and green consumers. Finally, opportunities for privately profitable “win–win” CSR are, in general, probably quite limited. Scores of studies have examined the link between CSR and corporate profits, and several recent meta-analyses conclude that on average the link, if it exists at all, is at best mildly positive, such that while CSR does not usually entail significant losses, neither does it generate significant profits. In other words, most CSR just pays for itself (Reinhardt et al. 2008; Margolis et al. 2007; Portney 2008). Another argument against promoting CSR in LAC is potentially even more damaging. As discussed below, the evidence that CSR actually results in better environmental management than otherwise would have occurred is very thin, and not because there has been no research on the topic. Moreover, some have argued that beyond being ineffective, CSR may do more environmental harm than good, particularly in developing countries, by “greenwashing” – creating a false impression that farms, firms, and society in general are making significant progress on environmental management, thereby diverting political, financial, and human resources from conventional regulatory activities that are likely to be effective. In other words, CSR, which has little or no additional environmental benefits, may substitute for regulatory action, which has greater benefits (Blackman and Sisto 2006; Blackman et al. 2013; Clapp 2005; Greer and Kenny 1996). For example, environmental regulators in Colombia devoted a significant share of their scarce political, financial, and human resources during the 1990s and 2000s to putting in place more than 60 voluntary agreements with polluting sectors (Blackman et al. 2013). By all accounts, the vast majority of these agreements failed to spur any activity at all, and recent evaluations conclude that even those reputed to be most successful resulted in few if any additional benefits. The opportunity cost of using scarce regulatory resources to promote CSR was more conventional regulation, which even if imperfect probably would have resulted in at least some additional benefits. 3.14.4 Evidence Evaluating the additional impact of CSR on environmental performance is challenging because it requires estimating a counterfactual – what environmental performance would



have been absent CSR. This challenge can be met by gathering data on environmental performance over time both for firms engaging in CSR and for similar control firms not engaged in it, and/or by developing an in-depth understanding of the firm-level decision making associated with CSR and environmental performance. Most empirical research on the impact of CSR concerns voluntary environmental programs and voluntary environmental agreements (versus unilateral actions) in industrialized countries. Although in some cases these programs and agreements may have significant environmental benefits, meta-analyses of research on both types of CSR suggest that, on average, they have very little, if any, additional effects. For example, Lyon and Maxwell (2008b: 723) conclude that “A growing body of empirical work suggests [voluntary environmental programs] generally have little impact on the behavior of their participants” (see also Sides and Darnall 2008; Koehler 2008). And with regard to voluntary agreements, OECD (2003: 14) finds “only a few cases where [voluntary] approaches have been found to contribute to environmental improvements significantly different from what would have happened anyway” (see also EEA 1997; Harrison 1999). Evidence from developing countries is thinner, but wholly consistent (Blackman 2008, 2010; Blackman et al. 2010, 2013). Evidence for whether CSR has environmental costs through greenwashing and diverting resources in developing countries is even thinner and confined to case studies (Blackman and Sisto 2006; Blackman et al. 2013; Clapp 2005; Laufer 2003).

3.15 Greening agriculture 3.15.1 Description Agriculture is the greatest threat to terrestrial biodiversity and ecosystem service provision in LAC. This threat operates through multiple channels (Grau and Aide 2008; IAASTD 2009; UNEP 2010; TNC 2005; Chomitz 2007; Tscharntke et al. 2012). First, agricultural extensification destroys and fragments habitat at the agricultural frontier and in mosaic lands, specifically remnant habitat patches, hedgerows, and trees. Second, agriculture degrades habitat by spurring extraction of timber and nontimber products in adjacent forests, introducing livestock and nonnative species (including invasives and GMOs), generating water pollution (from agrochemical use, processing plants, and sediment run-off), overexploiting surface water and groundwater, and in general disrupting natural processes. Third, agriculture contributes directly and indirectly to global climate change. It contributes 10–12 percent of the world’s total anthropogenic emissions of greenhouse gases, mostly methane and nitrous oxide, which are particularly potent greenhouse gases (Smith et al. 2007). In addition, agricultural extensification is a leading cause of forest clearing and degradation, which contribute an estimated 7–15 percent of global anthropogenic carbon dioxide emissions (Harris et al. 2012; Pan et al. 2011; van der Werf et al. 2009). Finally, agriculture can cause soil erosion, particularly on steeply sloped land and degraded pastureland. Erosion damages biodiversity and ecosystems services directly and can also reduce agricultural productivity, leading to further extensification. Of course, the specific effects of agriculture on biodiversity and ecosystem services depend on a host of site-specific factors, including the type of agriculture, the production methods used, and the geophysical context. That said, the broad environmental objectives of green agriculture are to prevent habitat loss from agricultural extensification and to minimize the environmental harms from agricultural intensification (TNC 2005; UNEP 2010).



3.15.2 Status and trends The status and trends in LAC agriculture and the relationship between extensive agriculture and forest area are discussed in Section 3.2. Here we briefly provide evidence on the growth of intensification. For example, in LAC, fertilizer use and fertilizer intensity (the amount applied per unit of land area) have both doubled since 1990 (Table 3.15-1; see Table A3.15-1 for country-level data). At the subregion level, fertilizer consumption in the Caribbean has actually declined by 65 percent over this time, although the intensity of fertilizer use was initially much higher than in Latin America (ECLAC 2011). 3.15.3 Issues The relationship between agricultural extensification and intensification is complex and affects agricultural policies aimed at biodiversity conservation (Lambin and Meyfroidt 2011; Angelsen and Kaimowitz 2001; Lee and Barrett 2001; Sayer and Cassman 2013). A common hypothesis is that without policies or a deficit of new land to block it, agricultural output is typically expanded via extensification. When such constraints bind, however, intensification results – that is, farmers seek to increase production on existing agricultural land. Although agricultural intensification has been suggested as a strategy for reducing pressure on natural systems by enabling global food needs to be met on a smaller land area, it can encourage extensification by increasing agricultural productivity and profitability. In addition, intensification can impinge on biodiversity and provisioning of local ecosystem services through increased use of agrochemicals, loss of remnant habitat patches, and the use of less biodiversity-friendly agricultural practices. And blocking extensification in one area can lead to extensification in another (i.e., leakage). Notwithstanding these complications, as noted above, a consensus has emerged that the main focus of efforts to stem loss of biodiversity and ecosystem services due to agriculture should be preventing extensification and reducing damage from intensification (TNC 2005; Chomitz 2007). Preventing extensification Strategies to reduce agriculture extensification include land-use planning, reform of perverse subsidies, better forest management, and strategic road planning to avoid opening access to new land (Chomitz 2007). Land-use planning can promote new agricultural production on lands where conservation would generate relatively few benefits – that is, on lands with relatively low levels of biodiversity and ecosystem services. Conflict between food Table 3.15-1 Fertilizer use and intensity, by subregion, 1990–2008 Region

Use (tonnes) 1990


Intensity (tonnes/ha) 2008

Percentage 1990 2000 2008 growth, 1990–2008

LAC 7,909,684 12,287,492 16,587,625 110 Latin America 7,182,090 11,983,780 16,329,388 127 Caribbean 727,594 303,712 258,237 –65 Source: ECLAC 2011.


17.3 23.3

Percentage growth, 1990–2008 101



production and biodiversity can be reduced by using the most productive areas for highyielding crops while leaving less productive areas for biodiversity protection, watershed protection, and carbon sequestration (Grau and Aide 2008). Reforming perverse economic incentives, including irrigation and agrochemical subsidies, could reduce rates of agricultural expansion. Better implementation of forest protection policies, including protected areas and comanagement, also can stem agricultural expansion. Land conversion taxes, payment for ecosystem services, and easements can be used for protecting habitat on private land. Finally, implementing systems to equitably allocate and enforce property rights can help block extensification. Reducing adverse effects of intensification A variety of best-management practices can help minimize the effects of agricultural extensification on biodiversity and ecosystem services (TNC 2005; Ongley 1996; McNeely and Scherr 2003; IAASTD 2009). Agrochemical use can be reduced by increasing the efficiency of application, replacing chemical fertilizers with natural ones, using integrated pest management, and maintaining diverse crop species and remnant habitat patches to support natural populations of pest enemies. Agricultural water pollution and sedimentation can be reduced by locating point sources of pollution away from water bodies, using physical and vegetative filters and barriers to prevent runoff, and again, increasing the efficiency of agrochemical use. Natural habitats can be maintained by locating fields and pastures in areas with appropriate soils, slope, and conditions while leaving less suitable and sensitive areas in natural habitat. Maintaining natural habitats within agricultural landscapes is particularly important because many imminently threatened species tend to be located in these types of productive landscapes. Strategies to increase adoption of best management practices include conducting training and information initiatives; reducing, decoupling, or eliminating agrochemical subsidies and/or taxing their use; abolishing policies that require use of intensive agricultural practices to gain access to land rights and crop insurance or that penalize tree growing; and promoting PES programs for carbon, biodiversity, water regulation, recreation, and pest control (UNDP 2010; Chomitz 2007; TNC 2005). NGOs, research institutions, and government extension programs are often underfunded for large-scale capacity building in sustainable agriculture. Thus, the knowledge of how to farm sustainably and with greater productivity is not reaching many producers (TNC 2005). Finally, agroforestry and silvopastoral systems, in which trees and shrubs are integrated with crops and livestock, can deliver many of the same ecological benefits as natural forests – including providing biodiversity habitat, sequestering carbon, facilitating aquifer recharge, and preventing soil erosion – albeit at lower levels. These systems comprise 200–357 million hectares in Latin America, including 14–26 million hectares in Central America and 88–315 million hectares in South America (Somarriba et al. 2012). Increased adoption of these mixed agricultural systems often requires providing training and information, linking farmers to markets, and securing land tenure (UNDP 2010; McNeely and Scherr 2003; Somarriba et al. 2012). Biofuels Biofuels have been advocated on the grounds that they reduce greenhouse gas emissions and fossil fuel dependence. However, biofuels may provide less energy than is expended (using



fossil fuels) in producing them (Pimentel and Patzek 2005; IAASTD 2009). In addition, as noted above, expansion of biofuel production sometimes has all the adverse environmental effects associated with both extensive and intensive agriculture. Monocropping, particularly of GMOs, in biofuel production threatens agricultural biodiversity. It is worth noting that the choice of biofuel species affects the type of land (e.g., soils and slope) that can be used and the likelihood of invasiveness. For example, oil palms can be grown on acidic soils, and some types of grasses can grow on steeper and drier soils (Grau and Aide 2008). 3.15.4 Effectiveness The appropriate strategies for minimizing the effects of agriculture on biodiversity and ecosystem services vary across space and land types, such that different approaches may be warranted. For example, different approaches may be relevant in highly fragmented landscapes, along the forest frontier, and in more remote areas (Chomitz 2007). Despite the complex interactions between intensification and extensification, a few LAC countries have simultaneously increased their forest cover and agricultural production, including Costa Rica, El Salvador, and Chile. The countries that have been successful have relied on various mixes of agricultural intensification, land-use zoning, forest protection, increased reliance on imported food and wood products, the creation of off-farm jobs, foreign capital investments, and remittances (Lambin and Meyfroidt 2011).

3.16 Targeting, data, and evaluation 3.16.1 Description Targeting of conservation investments prioritizes projects and locations to get the highest net benefits – the biggest “bang for the buck.” It can provide much greater biological benefits, and at lower costs, than less systematic but commonly used approaches, such as professional judgment or targeting based on only biological criteria (Naidoo et al. 2006; Boyd et al. 2012). It is critically important because financial, human, and technical resources available for conservation are scarce and interventions can be applied to only a limited number of areas. For example, Newbold and Siikamäki (2009) find that if conservation targeting is done optimally, spending 10 percent of the cost of restoring all upstream watersheds for salmon protection yields nearly 80 percent of the maximum possible ecological benefits achievable by protecting all watersheds. Similar findings are available elsewhere in the literature (e.g., Ferraro 2003; Naidoo et al. 2006). Though relative benefits from improved targeting will vary by application, it is reasonable to believe that they are generally large. However, information on the ecological benefits and economic costs of conservation projects across space is needed for targeting. The necessary ecological information ranges from basic data on species presence and the threats they face to more complex information about how biodiversity and ecosystem services respond to specific environmental changes and how conservation actions do and do not spur such changes. Economic information includes estimates of the costs of conservation options, including opportunity costs, and information on the economic value of the relevant species, biodiversity, and ecosystem services. Although targeting of conservation investments generally draws from ex ante information, conservation effectiveness can also be improved by ex post evaluations, which can facilitate improvements in both current and future conservation interventions.



3.16.2 Status and trends Despite considerable improvements over the past decade or so, the lack of information on species, biodiversity, and ecosystem services represents a significant constraint to conservation planning and evaluation. Even the most basic knowledge on species has significant gaps. For example, simple data on the number of species in taxonomic groups is limited outside some of the most studied groups, such as mammals and birds. Moreover, even for the best-known species and species groups, high-quality data on species abundance and its ecological drivers at different locations is lacking. Information on ecosystem services is yet more limited; their magnitude and economic importance is generally poorly understood beyond boilerplate lists of the types of ecosystem services potentially present at different locations. Also needed for targeting is more information on ecological production functions – how various ecosystems generate hydrological services, biodiversity habitat, etc. And whereas rigorous evaluation of economic development policies is becoming more and more widespread, this trend has only very recently begun in the conservation policy world (Miteva et al. 2012; Ferraro and Pattanayak 2006; Pullin and Knight 2009). The problem of poor biodiversity data and conservation policy evaluation is not specific to LAC – it is a worldwide challenge. However, LAC offers some notable examples of significant investments in improving the systematic collection and dissemination of biodiversity data. Leading LAC biodiversity data institutions include CONABIO in Mexico and INBio in Costa Rica, and leading international organizations active in LAC include DIVERSITAS and the International Council on Science (ICSU) (Larigauderie et al. 2012). The state of biodiversity knowledge, science, and research needs in LAC is exhaustively summarized and described in ICSU-LAC (2010). 3.16.3 Issues Enormous heterogeneity exists in LAC both in biodiversity and in what we know about it. Information is particularly limited for freshwater and marine systems (ICSU-LAC 2010; Revenga et al. 2000; Burke et al. 2001). In compiling this book, we identified the following specific needs, which are supported by existing biodiversity science assessments (ICSULAC 2010; Revenga et al. 2000; Burke et al. 2001; Matthews et al. 2000): •

Improved biodiversity inventories, including georeferenced species distribution information, especially for freshwater, coastal, and marine systems. Even in terrestrial systems, for which knowledge is greatest, much basic information is absent. For example, Brazil is one of two countries in LAC with published information on the number of described species; it is believed to host a total of 1.8 million species, only 170,000– 210,000 of which have been described (ICSU-LAC 2010). Better information on anticipated regional climate change patterns and biodiversity responses. Research on this topic at an ecosystem level is generally hindered by the lack of long-term data sets and regional data sets, with some notable exceptions. Improved understanding of and data on current and anticipated threats to biodiversity, including the distribution and drivers of threats. For example, there is significant need for improved geospatial information on past, present, and future land-use change, including specific causes and locations of land conversion and degradation, existing and planned hydro-development projects, road networks, coastal development, and marine habitat loss. Improved data on inland and marine fisheries harvest and status, water


• • • •

• •


usage and capacity data at the watershed scale, and biological and chemical water quality monitoring data are also needed. Finally, policy makers need a better understanding of tradeoffs between agricultural intensification and extensification at local scales (Sayer and Cassman 2013). Information on the distribution and impacts of invasive species, including development of early detection systems (Gardener et al. 2012). Standardized land-use and land-cover change monitoring systems and data. Improved understanding of ecosystem service production, including how values and threats vary across space and across managed versus unmanaged systems (Balvanera et al. 2012). Studies assigning monetary values to non-market environmental and ecosystem amenities and services. Although the number of such studies has been increasing, environmental valuation in LAC remains in its infancy. Ecosystem services and their values are location-based, which further stresses the need for improving understanding. Development of standardized biodiversity indicators to allow for comparison within and among countries (OECD 2012). Improved and centralized information on existing environmental policies and programs, including nonenvironmental sector policies that significantly affect the environment, as well as levels of enforcement. International sharing and coordination of data and research. With a few notable exceptions (e.g., CONABIO, INBio), a general problem is a lack of willingness of institutions to make data available online (ICSU-LAC 2010). The Global Biological Information Facility is an example of an information-sharing initiative (OECD 2012). The region would also benefit from consolidation of a network of ecological observatories to undertake research and long-term monitoring of how climate and land-use changes alter biodiversity (ICSU-LAC 2010).

3.17 Reduced emissions from deforestation and degradation 3.17.1 Description Global deforestation, most of which occurs in developing countries, accounts for 7–15 percent of total anthropogenic greenhouse gas emissions – roughly the same share as the transportation sector (Harris et al. 2012; Pan et al. 2011; van der Werf 2009; IPCC 2007). Although some have expressed skepticism (Blackman 2010), a number of studies suggest that the cost of abating these emissions compares quite favorably to the costs of cutting emissions in other sectors (Lubowski and Rose 2013; Kinderman et al. 2008; Naburs et al. 2007). As a result, international climate mitigation policy has increasingly focused on reducing emissions from deforestation and forest degradation (REDD) in developing countries. The acronym REDD has come to refer to a broad set of policies wherein the global community, typically led by industrialized countries, rewards (pays) developing countries for additional forest conservation – that is, conservation over and above the level that would have occurred absent rewards. Although climate mitigation is the main goal, this conservation would clearly have a variety of “co-benefits,” not the least of which would be preserving biodiversity. The term REDD+ is commonly used to refer to modifying REDD activities to prioritize some of these co-benefits. As discussed below, despite considerable effort, the international institutional framework for REDD has yet to coalesce. The lion’s share of the debate and negotiation about this framework has focused on the rules and procedures needed to include REDD in new



international treaties called protocols developed under the authority of the United Nations Framework Convention on Climate Change (UNFCCC), the 1992 treaty aimed at stabilizing greenhouse gases. At the end of the day, no matter how this framework is structured, REDD will boil down to creating new incentives for land managers in developing countries to implement or strengthen the same forest conservation activities discussed in other chapters of this book: protected areas, comanagement, land-use planning, subsidy reform, payments for environmental services, eco-certification, forest governance, and green agriculture. Because the institutional framework for REDD is still evolving, we structure our discussion of REDD slightly differently from that of other biodiversity policies. We first provide a brief overview of the main issues in the ongoing debate about REDD, then we summarize the history of efforts to implement REDD, and finally, we briefly review evidence on the effectiveness of REDD activities. 3.17.2 Issues6 Scope The scope of REDD could be subnational, national, or both. Subnational activities are projects that focus on preserving forests in a defined geographic area (e.g., via protected areas or forest certification), while national ones target an entire country (e.g., via subsidy reform and improved forest governance). Each has advantages (Myers Madeira 2008; Angelsen 2008b). Subnational activities are easier to implement and monitor. Moreover, they accommodate differences in institutional capacity across regions within a country. Therefore, by relying on subnational activities, it is possible to push ahead with REDD even when national capacity is limited. However, compared to subnational activities, national ones better control for leakage (i.e., the shifting of forest cover change from areas targeted by subnational policies to areas that are not targeted), make it easier to integrate REDD into national development policies, facilitate consistent monitoring within a country, and bolster country ownership of REDD. A “hybrid” or “nested” approach combines the two levels of activities: the idea is to begin with subnational activities and then graduate to national ones. The difficulty, of course, is to harmonize the two approaches. Reference levels As noted above, REDD rewards land managers for forest conservation that is additional to the level that would have occurred absent rewards. Defining this “reference” or “baseline” level is challenging. One option is to use historical rates of deforestation and degradation. But then countries with relatively low historical rates will need to cut deforestation and degradation even further to earn REDD payments. Stakeholders have argued that this unfairly punishes countries that have already taken action to conserve forests and rewards those that have not. In addition, using historical rates of deforestation and degradation to set reference levels will misrepresent the rate that would have prevailed absent REDD when recent developments – say changes in forest policy or international prices of timber or agricultural commodities – significantly raise or lower deforestation and degradation. A similar concern arises in any standard one-size-fits-all protocol for setting reference levels – individual circumstances may invalidate the result. The counterargument is that allowing entities to tailor reference levels to their own circumstances will encourage them to artificially inflate these levels. Typically, pilot REDD activities have carved out a middle ground, using historical rates of deforestation and



degradation as a starting point and making adjustments for national circumstances (Angelsen 2008a). Regardless of how the reference level is set, there are bound to be tradeoffs between “additionality” – the emissions reductions over and above those that would have occurred absent REDD – and participation. That is, relatively stringent (low) reference levels ensure additionality but discourage participation and vice versa. Monitoring Once a reference level has been established, REDD depends critically on accurate monitoring of deforestation and degradation relative to that level. Advances in remote sensing – data collection from satellites and airplanes – have made it increasingly possible to accurately measure deforestation (Wertz-Kanounnikoff and Verchot 2008). Measuring forest degradation is more problematic. Typically, monitoring entails three steps: (1) developing an inventory that describes the extent, types, and condition of forests, (2) measuring changes in these forests using a combination of remote sensing and field measurements, and (3) translating changes in forest cover to changes in carbon (Myers Madeira 2008). Standard protocols for monitoring, reporting, and verification (MRV) are still under development. An important consideration has been findings ways to accommodate differences in institutional and technological capacity across developing countries using a phased approach wherein countries would graduate from less to more sophisticated MRV methods. As with reference levels, there are bound to be tradeoffs between stringency and participation. Leakage Leakage occurs when REDD spurs increased deforestation or degradation outside of targeted areas. A number of mechanisms may be involved. For example, agents responsible for deforestation – including households, logging enterprises, and farms – may strategically shift their activities from targeted to nontargeted areas. Alternatively, or in addition, by restricting the supply of timber and agricultural goods in targeted areas, REDD activities can push up prices in nontargeted areas, which in turn boosts forest cover change. A number of strategies for controlling leakage have been suggested, including monitoring deforestation and degradation outside of targeted areas, targeting large areas or entire countries instead of limited areas, and neutralizing leakage with complementary REDD activities (Wunder 2008). Permanence Forests are inherently impermanent. Even leaving aside the effects of human activity, they can be adversely affected by fires and pests. As a result, paying land managers for REDD entails a risk that targeted forests will eventually disappear. A number of mechanisms have been proposed to mitigate this risk, including requiring a percentage of payments to land managers be held in reserve accounts, making intermittent REDD rewards conditional on continued conservation, and requiring land managers to acquire commercial insurance (Dutschke and Angelsen 2008). Capacity An omnipresent constraint in REDD policy is that, as discussed in foregoing sections of this book, many developing countries lack capacity to implement effective, efficient, and



equitable forest conservation policies (Kanowski et al. 2011). Efforts to accommodate this reality have focused on developing a three-phase approach for countries seeking to participate in REDD (Meridian Institute 2009). In the first phase, countries develop a national REDD strategy including proposed REDD activities, reference levels, and an MRV system. In the second stage, they implement these policies and measures. Only in the third phase are they paid for reducing emissions from deforestation and degradation below an agreed-upon reference level. Co-benefits and safeguards Aside from reducing greenhouse gas emissions, REDD could have additional positive and negative effects depending on the specific policy and the context (Brown et al. 2008; Phelps et al. 2010). As for the positive effects, REDD could lead to increases of all the ecosystem services that forests provide, including conserving biodiversity, facilitating aquifer recharge, and regulating surface water flows. In addition, depending on how the rewards are distributed, REDD could help alleviate poverty. But critics have argued that REDD could also exacerbate corruption, disenfranchise indigenous peoples and local communities that depend on goods and services provided by forests, and lead to substitution of industrial forest plantations for natural forests (Brown 2013). Most stakeholders agree it is important to put in place institutions and rules to encourage the positive effects and “safeguards” to discourage the negative ones. Financing Financing is perhaps the most pressing and contentious REDD issue. Financing is needed both for investments in REDD readiness – that is, preparing countries to participate in REDD – as well as to cover the costs of implementing national policies and compensating forest owners for the foregone profits from deforestation and degradation. There are two main financing options, which most stakeholders agree should be complements rather than substitutes (Dutschke and Wertz-Kanounnikoff 2008). The first is direct government-to-government funding. Already, as discussed below, overseas development assistance (ODA), such as the Norwegian government’s International Forests and Climate Initiative, comprises a significant portion of existing REDD funding. The second option is to rely on revenue generated by a market-based mechanism. The most widely discussed variant is a system wherein entities obligated to reduce their greenhouse gas emissions under the Kyoto Protocol or its successor meet those obligations by purchasing offsets generated by REDD. Taxes and auctioned emissions allowances also have been discussed. 3.17.3 Status and trends Progress under UNFCC As noted above, the lion’s share of the extensive debate about REDD in the past decade has been devoted to working out rules and procedures needed to include it in new UNFCCC protocols. The main venue for this debate has been annual UNFCCC conference of the parties’ (COPs) wherein parties to the UNFCCC (signatory countries) negotiate new protocols. What follows is a brief summary of important milestones in that debate. Under the 1997 Kyoto Protocol, Annex I Parties (i.e., those from industrialized and certain transitioning countries) agreed to emissions reduction targets, which could be met by



either reducing emissions, or purchasing credits generated by reductions in other countries, including developing countries. REDD was considered as a potential means of generating such credits, but the idea was ultimately abandoned because of concerns about several of the issues discussed above, namely monitoring, reference levels, permanence, and leakage. As a result, REDD credits have thus far been excluded from conventional market for carbon credits used by Annex I countries seeking to meet their Kyoto emissions reductions targets. At the 2005 COP-11 in Montreal, the Coalition of Rainforest Nations, led by Costa Rica and Papua New Guinea, reintroduced the idea of allowing REDD credits to be used as offsets under the UNFCCC. Their proposal was taken up in subsequent COPs and was ultimately included as part of the 2010 COP-16 Cancun agreements, albeit in broad terms without the level of detail needed for actual implementation under the UNFCCC. The agreement set a goal of national-level scope (although subnational activities could be used in transition to national ones), and encouraged parties to contribute to mitigation actions in the forest sector by (1) reducing emissions from deforestation, (2) reducing emissions from forest degradation, (3) conserving forest carbon stocks, (4) promoting sustainable forest management, and (5) enhancing carbon stocks. (Because they expand activities beyond REDD, the last three items led to the term REDD+.) In addition, the agreement codified the three-phase approach to participation described in Section Thorny somewhat technical issues of reference levels, MRV, safeguards, and financing were deferred to subsequent COPs. Unfortunately, since COP-16 in 2010, progress on settling those issues has been slow. Despite a series of deadlines, as of the 2012 COP-18 in Doha, all had yet to be resolved (Morgan 2012; Tuttle 2012). Progress outside of the UNFCCC Even as negotiations for including REDD in the UNFCCC continue, considerable progress has been made outside of that framework. Both multilateral and bilateral international cooperation agencies and coalitions have funded REDD readiness (including the development of national REDD strategies, MRV programs, and reference levels), pilot projects, and the development of a market for trading REDD credits. The three major multilateral schemes are the United Nation’s UN-REDD program, which has pledged $119 million to support development of national REDD strategies; the World Bank’s Forest Carbon Partnership Facility (FCPF), which has committed or pledged $240 million for various readiness activities and $219 million for a REDD credit-trading scheme; and the Forest Investment Program (FIP), a program supported by a consortium of multilateral organizations led by the International Bank for Reconstruction and Development, which has pledged $639 million for readiness and pilot activities (Peters-Stanley et al. 2012). Seventeen LAC countries are participating in these schemes (Table 3.17-1). In addition to these multilateral activities, bilateral agencies are providing significant funding for REDD readiness. For example, USAID has slated tens of millions of dollars for REDD readiness in Latin America. Total financing from all multilateral and bilateral organizations exceeds $2.4 billion (Table 3.17-2). Aside from multilateral and bilateral funding, the voluntary carbon market represents another source of REDD financing. While credits from REDD are excluded from conventional markets for carbon credits used by Kyoto Protocol Annex I countries, a relatively small voluntary market for REDD credits generated by subnational pilot projects has emerged outside of the UNFCCC. Various organizations, including Voluntary Carbon Standard, Climate Community, and Biodiversity Alliance Standard verify and certify these



Table 3.17-1 LAC countries participating in major multilateral REDD readiness schemes Country Argentina Bolivia Brazil Chile Colombia Costa Rica Ecuador El Salvador Guatemala Guyana Honduras Mexico Nicaragua Panama Paraguay Peru Surinam




9 9


9 9

9 9 9 9 9 9 9 9 9 9 9 9 9





Note: FCPF = World Bank Forest Carbon partnership Facility FIP = Forest Investment Program Source: Westholm et al. 2012.

REDD credits, which are then sold (at a discounted price relative to the official market) to buyers seeking to reduce emissions for reasons other than Kyoto obligations. Latin America leads the developing world in the supply of REDD credits to this market. In 2011, credits generated by 70 pilot projects in Latin America representing a total forest area of 956 million hectares and 7.7 million tons of carbon were traded on the voluntary market (Peters-Stanley et al. 2012; Hall 2012). Finally, although REDD is not yet fully functional under the UNFCCC and no comprehensive international carbon market exists, several regional and state-level programs in Europe and the United States may provide opportunities for the inclusion of REDD, although no large-scale implementation has yet emerged. The largest and most comprehensive cap-and-trade system is the European Union Emissions Trading System (EU ETS), which has been in operation since 2005. It accepts offsets in the form of Clean Development Mechanism (CDM) or Joint Implementation (JI) credits, excluding land use and nuclear power. In 2008, the European Commission investigated including REDD offsets in the third phase of EU ETS but had serious reservations about several of the same issues that have concerned UNFCCC negotiators, including reliability of MRV systems and the permanence of credits (Commission of the European Communities 2008). As a result, land-use-based credits will not be included in the EU ETS until possibly after 2020. In the United States, the subnational-scale program with the most potential for REDD is California’s Global Warming Solutions Act, also known as Assembly Bill (AB) 32 (California Legislature 2006). The act aims to reduce California’s emissions to 1990 levels by 2020 through a combination of regulations and cap-and-trade markets. Offsets are included in the market, and a number of offset design methodologies, including one for avoided deforestation, are being developed for the system.



Table 3.17-2 Total REDD financing reported by funders, by country, 2013 Country

Reported by funders ($ millions)

Argentina Belize Bolivia Brazil Chile Colombia Costa Rica Cuba Dominican Republic Dominica Ecuador El Salvador Guatemala Guyana Haiti Honduras Jamaica Mexico Nicaragua Panama Paraguay Peru Surinam Venezuela Total

1.12 0.1 21.96 561.13 0.04 5.12 60.27 0.15 1.8 1.7 63.89 0.1 7.01 83.32 4.19 0.15 5.7 1,537.35 2.2 3.69 11.3 58.47 0.42 0.69 2,431.87

Note *Voluntary unverified reports. Source: REDD+ Database 2013.

3.17.4 Effectiveness Given that rigorous evaluation of all manner of well-established conservation policies is rare and that REDD policy remains nascent, it is not surprising that the evidence base for the effectiveness of REDD in stemming forest-cover change and conserving biodiversity is quite thin. Although subnational REDD pilot projects – which as just noted are rapidly proliferating in LAC – provide fodder for evaluation, thus far, serious evaluation of these projects has been limited. Caplow et al. (2011) review the published and gray literature, evaluating 20 “pre-REDD” projects (defined as projects focusing primarily on REDD in a non-Annex I country, having been launched after COP-1 but before COP-13, and having estimated their impacts on net greenhouse gas emissions), 15 of which are in Latin America. The authors found only five evaluations of these projects, which generally lacked rigor. They conclude that drawing lessons from this evidence base is “tenuous.”

Notes 1 The 39 countries, by region in descending rank in terms of forest area, are [Latin America: Brazil, Peru, Mexico, Colombia, Bolivia, Venezuela, Argentina, Guyana, Suriname, Ecuador, Honduras] [Asia: China, Australia, Indonesia, India, Myanmar, Papua New Guinea, Malaysia, Thailand,


2 3

4 5



Cambodia] [Africa: the Democratic Republic of Congo, Sudan, Angola, Zambia, Tanzania, Central African Republic, Congo, Gabon, Cameroon, Mozambique, Mali, Chad, Nigeria, Cote d’Ivoire, Senegal, Burkina Faso, Niger, Gambia, and Togo]. At the country level, empirical research focused on forests has consistently linked insecure property rights to deforestation (Deacon 1999; Bohn and Deacon 2000; Culas 2007). It is important to mention two caveats. First, the program has had little impact on deforestation partly because there has been little deforestation. As Pagiola (2008) notes, it is difficult to disentangle the effect of the program from other factors that have contributed to reduced deforestation, including strict command-and-control regulations and declining returns to pasture. Second, most rigorous statistical analyses of the program have deemphasized or ignored the program’s effects on reforestation and on subnational effects, which according to some scholars, have been significant (Daniels et al. 2010; Morse et al. 2009). This section draws on Sedjo and Siikamäki (2009) and Siikamäki and Chow (2008). The World Bank has also undertaken its own effort to develop estimates of “comprehensive wealth accounts” for more than 150 developing and developed countries (World Bank 2011). Its estimates suggest that natural resources make up 21–35 per cent of total wealth in developing countries. It has also estimated the value of protected areas as a share of GDP, finding they range from 23 to 1 percent of GDP (Hamilton 2013). This section draws heavily upon Angelsen (2008b) and Myers Madeira (2008).


Lines of action

As discussed in Chapter 1, one objective of this book is to recommend five “lines of action” for conserving biodiversity in LAC. That is a challenging task. As we have seen, LAC is exceptionally diverse in terms of ecosystems, politics, culture, and economics. A wide array of site-specific factors determine what policies are most appropriate in any given place, and some policies may be effective in some places but not in others. Two caveats follow from those observations. First, our recommendations must necessarily be somewhat broad-brush and general. We have attempted to make them specific enough to be useful without being so specific that they do not apply broadly. Second, brief mention or even omission of a policy from our list does not necessarily imply that the idea is not worth pursuing in some instances. In general, based on the evidence and analysis in Chapters 2 and 3, we have prioritized policies meeting at least one – and in many cases several – of the following criteria: • • • •

the policy specifically addresses a pressing threat; and/or has been proven effective, and ideally cost-effective, in at least some situations; and/or leverages past investments in conservation; and/or helps address pervasive institutional obstacles to effective conservation.

Table 4.1 indicates which criteria were paramount in selecting each line of action. We further explain our reasons for selecting each line of action in the subsections that follow. Each of the next five subsections describes a line of action. The last subsection explains why we have not placed more emphasis on several of the policies discussed in Chapter 3, namely eco-certification, ecotourism, bioprospecting, and CSR programs. Each of the subsections describes a line of action, in three parts. The first explains our reasons for selecting the line of action, and the second gives specific recommendations. In parentheses, we refer readers to earlier sections of the book that provide evidence supporting the rationale and recommendations. The third part briefly describes how the overall line of action – and each component intervention – is expected to affect biodiversity and suggests indicators to measure those effects. A few explanatory remarks about the indicators are in order. We propose indicators of the effectiveness both of the broad line of action and of the specific recommendations. In some cases, the same indicator applies to both. For each recommendation, indicators are listed in order of coarse to fine. For expositional clarity, we describe each indicator as a level at a given point in time. However, we are really interested in changes in these levels, which would obviously require repeated measurement at different points in time. For example, to measure the effectiveness of our recommendation to promote green agricultural practices,


Lines of action

Table 4.1 Principal criteria met by each line of action Criterion

(1) Promotes (2) Strengthens (3) Improves green terrestrial environmental agriculture protected areas governance and comanagement

(4) Strengthens coastal and marine resource management

Addresses pressing threat Has been proven effective Leverages past investments Addresses pervasive obstacles to conservation




(5) Improves biodiversity data and policy evaluation

9 9 9


we propose measuring tons of chemical fertilizer used in a given year, but repeated measurements are required to track changes in fertilizer use over time. We specify the units for each indicator in brackets.

4.1 Green agriculture 4.1.1 Rationale This line of action addresses a pressing threat to LAC biodiversity (Table 4.1). Reducing the adverse effects of agriculture on biodiversity and ecosystem services in LAC is critically important for three reasons: 1

2 3

Agriculture is the most important driver of terrestrial biodiversity loss in LAC. To be precise, it is the top driver of terrestrial habitat loss and degradation, which in turn is the preeminent threat to terrestrial biodiversity. Agriculture destroys and degrades terrestrial habitat through two broad channels: extensification in frontier areas and mosaic lands, largely through expansion of export-oriented agroindustry; and intensification characterized by heavy use of water, soils, agrochemicals, and GMOs (Sections 3.2, 3.13). Most of the imminently threatened terrestrial species in LAC live in mosaic lands used for agriculture, not in remote contiguous landscapes (Section 3.2). Pollution, forest clearing, and forest degradation linked to agricultural extensification and intensification in LAC also destroy and degrade freshwater and marine biodiversity and ecosystem services (Sections 2.1, 2.2, 3.5).

4.1.2 Recommendations Conservation of biodiversity and ecosystem services in LAC can be advanced by stemming agricultural extensification and mitigating the adverse effects of agricultural intensification. We refer to these two broad activities as greening agriculture. The following actions would spur it:

Lines of action •


Reform agricultural subsidies. Subsidies for agricultural outputs (crop price supports) and inputs, including irrigation water, agrochemicals, land, and infrastructure, create economic incentives for agricultural extensification and environmentally unfriendly activities associated with agricultural intensification. Reducing or eliminating subsidies, therefore, can mitigate these adverse effects. Political challenges of subsidy reform sometimes can be addressed by decoupling subsidies – that is, returning their value to producers as lump sums unconnected to specific goods or activities (Sections 3.7, 3.15). Strengthen land-use planning. In principle, land-use planning (in addition to protected areas and comanagement, discussed below) can stem both agricultural extensification and certain types of particularly worrisome intensification. However, in LAC it is often undermined by failure to prioritize environmental goals among other aims (e.g., extending infrastructure and reducing disaster risks) and by weak implementation, monitoring, and enforcement. It can be strengthened by filling gaps and addressing inconsistencies in written regulations for land-use planning, explicitly incorporating environmental goals into planning regulations, building capacity in appropriate institutions, helping to enhance coordination of the myriad governmental institutions typically involved in land-use planning, and promoting experiments with habitat offsets (Sections 3.3, 3.6, 3.12). Promote green agricultural practices. A wide variety of green agricultural practices can help mitigate environmentally harmful effects of intensive agriculture in LAC, including natural fertilizers, pest and weed management, vegetative barriers, hedgerows, ground cover, terracing, agroforestry, silvopastoral systems, and more precise application of agrochemicals. Strategies to speed dissemination of such practices include conducting training and information initiatives and reforming subsidies that create disincentives for their use. Promoting research and development – for example, through the Consultative Group for International Agricultural Research centers – can also encourage green technological change (Sections 3.7, 3.8, 3.13). Improve payments for ecological services (PES) initiatives. PES can help stem agricultural extensification and create economic incentives to use green agricultural practices, particularly in agricultural mosaic lands where biodiversity is most threatened and private ownership is more likely. LAC leads the world in PES implementation, yet considerable evidence indicates that many PES initiatives lack design and implementation elements widely considered prerequisites for effectiveness, such as strictly enforced conditionality and targeting for ecosystem service provision and additionality. This is particularly true of the large-scale national programs in Brazil, Costa Rica, Ecuador, and Mexico that, arguably, could have the biggest impact on biodiversity. Hence, green agriculture can be advanced by initiatives aimed at improving PES design and implementation both for existing PES programs, especially large ones, and for planned programs.

4.1.3 Expected benefits and indicators OVERALL EFFECTIVENESS

In general, we expect this line of action to help conserve biodiversity by reducing agricultural extensification and mitigating the effects of agricultural intensification, which in turn


Lines of action

would stem habitat loss and degradation. The overall effect of the line of action through agricultural extensification could be measured by changes in • •

agricultural land area [ha]; and natural habitat land area, by habitat type (forest, grassland, desert, wetland, etc.) [ha].

The overall effect of the line of action through mitigating effects of agricultural intensification could be measured by changes in • • • •

agrochemical use (both total use and intensity) [tons, tons/ha], including fertilizers and pesticides; species richness in agricultural landscapes [no. species/ha]; surface water quality [biochemical oxygen demand and total suspended solids]; soil erosion, including – soil area [m2] – soil depth [m] – soil volume [m3]; and the percentage of crops that are traditional varieties [%].


Reforming agricultural subsidies can help conserve biodiversity by dampening economic incentives for agricultural extensification and intensification, which destroy and degrade habitat. The effectiveness of efforts to reform agricultural subsidies can be measured by changes in • • • •

the nominal rate of assistance in the agricultural sector (government subsidy as a percentage of gross value of agricultural production) [%]; gross agricultural subsidies [$]; the nominal rate of assistance in the agricultural sector counting only subsidies that spur environmental and natural resource degradation [%]; and gross agricultural subsidy, counting only subsidies that specifically spur environmental and natural resource degradation [$].


Strengthening land-use planning can help conserve biodiversity by stemming agricultural extensification, which destroys and degrades habitat. Simple quantitative indicators of the strength of land-use planning are difficult to identify. Therefore, we propose •

national-level qualitative expert assessment of changes in – the extent to which legal and regulatory framework for land-use planning is clear, complete, and internally consistent [qualitative]; – the extent to which a feasible mechanism for coordinating land-use planning activities across various institutions exists [qualitative]; – existence of actual land-use plans [qualitative]; – emphasis on environmental goals – specifically, stemming habitat loss and fragmentation – in land-use plans [qualitative]; and – monitoring and enforcement of land use plans [qualitative].

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Promoting green agricultural practices can help conserve biodiversity by mitigating the harms of agricultural intensification, which destroys and degrades habitat. The effectiveness of efforts to promote green agricultural practices can be measured by •

the percentage of farms or farmland using selected environmentally friendly agricultural practices, including [%]: – natural fertilizer – natural pest and weed control – vegetative barriers – ground cover – terracing – deviation canals and infiltration holes – agroforestry – silvopastoral systems; and agrochemical use (both total use and intensity), including fertilizers and pesticides [tons and tons/ha].


Improving PES initiatives can help conserve biodiversity by stemming agricultural extensification and mitigating the harms of agricultural intensification, both of which destroy and degrade habitat. The effectiveness of efforts to improve PES can be measured by •

program-level qualitative expert assessments of changes in PES program design and implementation aimed at enhancing targeting for additionality and ecosystem services provision, and enforcement of conditionality [qualitative]; and rigorous statistical analyses of PES effects on land-cover change using remote sensing data to measure land-cover change and statistical techniques to control for nonrandom siting (as described in Blackman 2013).

4.2 Strengthen terrestrial protected areas and comanagement 4.2.1 Rationale This line of action addresses a pressing threat to LAC biodiversity, entails a mechanism (protected areas) that has been proven effective in some circumstances, and leverages past investments in conservation (Table 4.1). More specifically: • •

PAs and comanagement have the potential to stem habitat loss, which is the preeminent threat to terrestrial and freshwater biodiversity (Section 2.1). Over the past 20 years, hundreds of new terrestrial PAs have been established in LAC such that today, one-fifth of the region’s land is protected, the highest share in the world (Section 3.1). Although some PAs are effective at conserving biodiversity and protecting ecosystem services, a large proportion are not particularly effective because they are fragmented, poorly managed, and underfunded (Section 3.1). Corridors among biodiversity habitats, including those in PAs, will serve as an important means of promoting adaptation to climate change (Section 3.1).


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Forest comanagement, including indigenous territories, extractive reserves, and community-managed concessions, has proliferated rapidly in LAC during the past 20 years. Today, it covers 22 percent of LAC forests (Section 3.2). Very little rigorous evidence indicates that forest comanagement is generally effective. Rather, its effectiveness depends critically on site-specific factors.

4.2.2 Recommendations The following actions would strengthen terrestrial PAs and forest comanagement: •

Build domestic capacity to sustainably finance existing PAs. According to a recent study, current allocations to LAC PAs, which amount to just over $1 per hectare per year, cover only about half of what is needed for minimally effective management, including developing management plans and having basic monitoring and enforcement. Planning, human capacity, and institutions are needed to build domestic capacity to sustainably finance existing PAs (Section 3.1). Consolidate and establish more corridors among existing PAs. Many LAC PAs are too small and fragmented to effectively stem the loss of biodiversity and ecosystem services. Large and/or uninterrupted areas, which help block access and discourage extraction, are required for some species and natural processes, and help facilitate climate adaptation. Hence, consolidating and establishing corridors among existing PAs – both within and between countries – can improve their effectiveness (Section 3.1). Selectively expand PA networks. Despite overall high levels of protection in LAC, many LAC areas fall far short of the widely accepted global goal of protecting 10 percent of all national territory and 10 percent of all biomes (Section 3.1). Expansion should take into account anticipated climate change. Provide additional support for improved comanagement. Local communities often do not have the requisite experience and resources to effectively and sustainably manage forests. Infrastructure, subsidies, technical assistance, and research can help meet this need (Section 3.2).

4.2.3 Expected benefits and indicators OVERALL EFFECTIVENESS

In general, we expect this line of action to help conserve biodiversity by stemming habitat loss and degradation. The overall effect of the line of action through agricultural extensification could be measured by • rigorous assessment of the effect of PAs and comanagement on forest cover using remote sensing data to measure land-cover change and statistical techniques to control for nonrandom siting (as per Blackman 2013). BUILD DOMESTIC CAPACITY TO SUSTAINABLY FINANCE EXISTING PAS

Building domestic capacity for sustainably financing existing PAs can help conserve biodiversity by increasing their effectiveness in stemming habitat loss and degradation. The success of efforts to build this capacity can be measured by

Lines of action • • •


changes in funding per hectare of protected area [$/ha]; the existence of a dedicated permanent budget for PAs (e.g., line item in annual budget) [0/1]; and expert assessment [qualitative].


Consolidating and establishing more corridors among existing PAs can help conserve biodiversity by blocking human access and thereby stemming the destruction and degradation of habitat, and it can improve habitat by facilitating natural processes and behaviors such as migration as well as climate adaptation. The effectiveness of efforts to implement this recommendation can be measured by: • • •

changes in fragmentation [edge/area ratio]; changes in connectivity [total area of connected PAs/total area]; and expert assessment of connectivity, taking into account the ecological value and effectiveness of corridors [qualitative].


Expanding PAs can help conserve biodiversity by stemming habitat loss and degradation. Expansion can be measured by changes in • •

the percentage of national territory under protection [%]; and the percentage of each of a country’s major biomes protected [%].


Improved comanagement can help conserve biodiversity by stemming habitat loss and degradation. Provision of additional support for improved comanagement can be measured by changes in • • •

funding for comanagement per hectare [$/ha]; the number of programs per hectare aimed at supporting comanagement [no./ha]; and the number of comanagement extension agents per hectare [no./ha].

4.3 Improve environmental governance 4.3.1 Rationale This line of action addresses pervasive institutional obstacles to effective conservation in LAC (Table 4.1). Environmental governance refers to the exercise of economic, political, and administrative authority by state institutions to manage the environment and natural resources. Improved environmental governance is needed in LAC for the following reasons: •

Implementation – including monitoring and enforcement – of environmental regulations and policies in LAC is weak. Most LAC countries have basic (if not fully fleshed out and internally consistent) environmental laws, regulations, and policies, but fail to


Lines of action

fully or consistently implement them. Implementation is critically important both to state-led conservation policies, such as PAs and land-use planning, and to incentivebased policies like bioprospecting and PES. Closely related drivers of weak implementation include a lack of political will and inadequate human, technical, and financial resources (Section 3.6). Government economic development and other policies in agriculture, fishing, transportation, tourism, and urban development have enormous consequences for biodiversity and ecosystem services. Yet in most cases, decision making linked to these policies does not adequately take into consideration potential effects on the environment in general and on biodiversity and ecosystem services more specifically (Sections 3.3, 3.4, 3.6, 3.7).

4.3.2 Recommendations The following actions would strengthen environmental governance in LAC: •

Build human capital. Knowledge about, and expertise in biodiversity conservation and ecosystem services is lacking in many LAC state institutions, particularly those not directly engaged in environmental management. Building such knowledge and expertise can enhance both the ability and the willingness of government institutions to protect the environment. Information campaigns, targeted training, revised academic curricula, and new requirements for hiring, promotion, and professional certification can further this goal (Section 3.6). Mainstream biodiversity. Mainstreaming biodiversity refers to incorporating biodiversity considerations into governance in state institutions–both those focused on environmental management and those not focused on it–that have environmental effects. Mainstreaming can be promoted by enhancing coordination between environmental management and economic development institutions (often sorely lacking), creating units in economic development institutions devoted to environmental management, strengthening requirements for environmental impact assessments, and building human capital through the means listed above (Section 3.6). Promote national environmental accounting. Environmental national accounting entails systematically collecting and disseminating data on environmental quality and natural resources and (ideally) incorporating them into national income measures. Although evidence of the benefit is lacking – partly because the practice is nascent in LAC – in principle, environmental accounting creates incentives to improve environmental governance by enhancing transparency, accountability, and data availability (Section 3.13).

4.3.3 Expected benefits and indicators OVERALL EFFECTIVENESS

In general, we expect this line of action to help conserve biodiversity by improving the effectiveness of a wide range of state-led conservation policies, including protected areas, command-and-control fisheries and forest regulations, and PES, that in turn address a wide range of threats to biodiversity, including habitat loss, pollution, and overexploitation. The overall effect of the line of action could be measured by

Lines of action • •


a national-level overall index of environmental management effectiveness, including the indexes being developed by IDB [index]; and indicators collected for national environmental accounting [various].


Building human capital in the manner described above can help conserve biodiversity by enhancing expertise in biodiversity conservation within government and raising awareness of biodiversity issues in civil society, which in turn improves stakeholders’ ability and willingness to undertake and implement a wide range of conservation policies. Enhanced expertise in government institutions can be measured by changes in • the percentage of staff with specialized training (degrees, certificates) in conservationand environment-related fields, by level of position, and by institutional focus of environmental management, transportation, agriculture, tourism, fisheries, and finance ministries [%]. Enhanced awareness in civil society can be measured by • • •

conservation and environmental management requirements in curricula at primary and secondary educational levels and postsecondary certificate and degree programs [0/1]; changes in government budget allocations for conservation and environmental management information campaigns [$]; changes in the number of people exposed to environmental awareness campaigns by outlet (newspaper, radio, internet, etc.) and geographic area [no. people per outlet per geographic area]; and changes in awareness of basic conservation and environmental management concepts, as measured by population surveys [depending on survey instrument].


Mainstreaming biodiversity in the manner described above can help conserve biodiversity by improving stakeholders’ ability and willingness to stem or mitigate threats to biodiversity (habitat loss, pollution, overexploitation, etc.) created by economic development policies, programs, and interventions. Simple quantitative indicators of mainstreaming are not easy to identify. Therefore, we propose measuring it by •

qualitative expert assessment, taking into account: – the existence and effectiveness of conservation and environmental management institutions in economic development agencies covering agriculture, fisheries, forestry, tourism, and transportation [qualitative]; – the extent to which conservation and environmental management are considered in economic development projects [qualitative]; – adequacy of the legal and regulatory framework for safeguarding against environmental damage from economic development projects (environmental impact assessments) [qualitative]; and – the quality and extent of coordination between environmental management and economic development institutions [qualitative].


Lines of action


National environmental accounting can help conserve biodiversity by enhancing transparency and accountability in conservation policy making, which in turn can help make such policy more effective. It can be measured by changes in the • •

• • •

existence of national environmental accounts [0/1]; completeness of national environmental accounts based on – the number of environmental and biodiversity stocks measured [no.] – the frequency of measurements [no./yr]; online availability of environmental accounts [0/1]; monetization of environmental accounts [0/1]; and incorporation of environmental accounts into national income measures (e.g., green GDP) [0/1].

4.4 Strengthen coastal and marine resource management 4.4.1 Rationale This line of action addresses a pressing threat to LAC biodiversity (Table 4.1). More specifically, strengthened coastal and marine resource management is critical for several reasons: •

LAC coastal ecosystems are extremely rich in biodiversity and provide valuable ecosystem services, including storm protection, filtration, and nursery and other types of habitat (Section 2.2). Human populations in LAC are heavily concentrated along the coasts and often depend directly or indirectly on economic activities in coastal areas, including fishing and tourism (Section 2.2). Both coastal systems and marine systems – notably fisheries – are among the most threatened in LAC, and threats to these ecosystems are expected to intensify because of population and economic growth and climate change (Section 2.2).

4.4.2 Recommendations The main threats to coastal and marine ecosystems and biodiversity are overexploitation, especially overfishing; pollution from agricultural, urban, and industrial sources; habitat loss and degradation, and climate change. We therefore propose the following actions: •

Reduce overexploitation of biological resources, especially overfishing. A twopronged approach is needed. For large commercial fisheries, the most effective approach is likely to be granting fishers individual or collective rights, and improving planning, enforcement, and monitoring, partly to facilitate rights-based regimes. For small artisanal fisheries, the portfolio of measures includes granting rights to fishers, comanagement, participatory planning and management, individual and community quotas, and coastal and marine protected areas. Comanagement in combination with marine PAs holds particular promise; it can improve governance and fisheries performance and reduce management costs. For both large and small fisheries, implementing

Lines of action


these policies entails raising public awareness and building capacity among regulators and fishers. Lack of public support and, therefore, political will often thwarts management efforts. Therefore, it is important to raise awareness among civil society and regulators (Section 3.4). Improve water pollution control from agricultural, urban, and industrial sources. Agricultural nonpoint source pollution is one of the most important pollution sources, and is addressed in the Green Agriculture line of action. Reducing urban wastewater pollution requires investment in municipal wastewater treatment technology and infrastructure. Pollution also can be reduced by improving monitoring and enforcement of existing regulations (Sections 3.5, 3.6, 3.15). Strengthen coastal planning and management. Coastal planning and management are the coastal analog of land-use planning. Like land-use planning, they require an ecosystem-based approach to integrating and coordinating policy interventions, including pollution prevention, urban development, fisheries management, and tourism management. Regulatory activities include zoning, access restrictions, monitoring and enforcement, and participatory planning and management (Sections 3.3, 3.6). Goals of coastal planning and management are (1) enhancing protection against damages due to economic development and climate change; and (2) mitigating unavoidable damages. Enhanced protection calls for both improving the management of current protected areas and establishing new protected areas. Marine ecosystems are complex and highly connected systems, so new protected areas require case-by-case assessment. For example, extending the size of protected areas may be less effective than connecting networks of smaller reserves. Mitigation involves first minimizing ecological damages by, for example, ensuring that urban development is not sited in highly threatened or sensitive areas and requiring best environmental management practices. A second line of defense is offsetting unavoidable damages using mitigation offsets. That approach is nascent in LAC and sometimes faces significant institutional obstacles, but its success in industrialized countries suggests it merits pilot applications and capacity-building efforts. Sealevel rise and ocean acidification due to climate change threaten coastal systems, so protection efforts need to facilitate adaptation (Sections 3.4, 3.5, 3.12).

4.4.3 Expected benefits and indicators OVERALL EFFECTIVENESS

In general, we expect this line of action to help conserve biodiversity by stemming overexploitation, pollution, and habitat destruction and degradation. Its overall effect could be measured by changes in: • •

the area of critical habitat, including mangroves, seagrasses, and coral reefs [ha]; and the threat levels of indicator marine species [various].


The benefit of reducing overexploitation of biological resources is obvious. Reduced overexploitation can be measured by changes in:


Lines of action

landings by species normalized by fishing effort, a proxy for robustness of the marine stock [tons/boat-meters per km2]; and the number and percentage of fish stocks classified as overexploited and fully exploited [no., %].


Improving water pollution control can help conserve biodiversity by reducing habitat destruction and degradation and (obviously) pollution. It can be measured by changes in: • •

the percentage of urban wastewater treated [%]; and surface water quality [biochemical oxygen demand and total suspended solids].


Strengthening coast planning and management can help conserve biodiversity by reducing pollution and habitat destruction and degradation. It can be measured by: • • • • •

changes in the percentage of national marine territory protected [%]; changes in the percentage of national coastal terrestrial area projected [%]; changes in the area of protected critical coastal habitat by type (mangroves, seagrasses, coastal wetlands, and coral reefs) [ha]; changes in the area of restored critical coastal habitat by type [ha]; and national-level expert assessment taking into account [qualitative]: – the extent to which legal and regulatory framework for coastal zone planning is clear, complete, and internally consistent [qualitative]; – feasible mechanisms for coordinating coastal zone planning activities across various institutions [qualitative]; – actual coastal zone plans [qualitative]; – the emphasis on environmental goals in coastal zone plans [qualitative]; – the extent to which climate change is considered in coastal zone plans [qualitative]; and – the level of monitoring and enforcement of coastal zone plans [qualitative].

4.5 Improve biodiversity data and policy evaluation 4.5.1 Rationale This line of action addresses a pervasive institutional obstacle to effective conservation in LAC (Table 4.1). Improving the collection and dissemination of data on biodiversity and the evaluation of biodiversity conservation policies is important for the following reasons: •

Better planning and evaluation of conservation policies are urgently needed. Specifically, systematic ex ante targeting of conservation policies and ex post evaluation of these policies are needed to ensure that they generate the most “bang for the buck.” That, in turn, is imperative because financial, human, and technical resources available for

Lines of action


conservation in LAC are scarce and interventions can be applied to only a limited number of areas. However, targeting and evaluation require data on the status of, trends in, and value of biodiversity and the expected ecological benefits and costs of policies across space (Section 3.16). Such data are, in general, sorely lacking in LAC. Even the most basic knowledge on species has significant gaps. For example, the number of species in taxonomic groups is unknown outside well-studied groups like mammals and birds. Although rigorous evaluation of economic development policies is becoming more and more widespread, this trend has only very recently penetrated the conservation policy world (Section 3.16).

4.5.2 Recommendations The following actions would strengthen biodiversity and biodiversity-related data collection and dissemination in LAC: •

• •

Collect and compile data on LAC biodiversity. Particular needs include biodiversity inventories, especially for freshwater, coastal, and marine systems; information on anticipated regional climate change patterns and biodiversity responses; data on current and anticipated threats to biodiversity; information on the distribution and effects of invasive species; studies assigning monetary values to nonmarket environmental and ecosystem amenities and services; standardized biodiversity indicators; standardized monitoring of land-use and land-cover change; and centralized information on existing environmental policies and programs (Section 3.16). Encourage coordinated online dissemination of biodiversity data in digital format (Section 3.16). Promote policy evaluation. Systematic, rigorous objective evaluation of conservation policies is needed, and dissemination of findings and their use in policy making should be incentivized. This can be accomplished by making evaluation a required component of policy planning and implementation at government agencies and of lending and grant making at foundations, bilateral, and multilateral international cooperation agencies; supporting training in monitoring and evaluation; and establishing regional centers of excellence for conservation policy evaluation and analysis (Section 3.16).

4.5.3 Expected benefits and indicators In general, we expect this line of action to help conserve biodiversity by improving the planning, targeting, and management – and therefore the effectiveness – of a wide range of public- and private-sector-led conservation policies on bioprospecting, land-use planning, fisheries and forest regulations, protected areas, and payments for environmental services. These policies in turn address such threats to biodiversity as climate change, habitat loss, pollution, and overexploitation. COLLECT AND COMPILE DATA ON LAC BIODIVERSITY

As discussed in the book, critical data on LAC biodiversity are sorely lacking, and systematically cataloging and prioritizing them would be an enormous task. Identifying indicators measuring specific data gaps is well beyond the scope of this book. Therefore, we recommend that the overall effect of the line of action could be measured by:


Lines of action

expert assessment of the conservation-related data gaps for: – biological data (e.g., species inventories, ranges and threats, biodiversity and ecosystem functions) [qualitative]; – socioeconomic data (e.g., cost of conservation, value of biodiversity and ecosystem services) [qualitative]; – policy effectiveness data based on rigorous quantitative evaluations [qualitative].


Clearly, the above-mentioned benefits of LAC biodiversity data depend critically on their availability, which can be measured by • • •

the existence of a centralized country-level web portal for biodiversity data dissemination [0/1]; the existence of a centralized regional-level web portal for biodiversity data dissemination [0/1]; and qualitative expert assessment taking into consideration the quality and comprehensiveness of web-based dissemination [qualitative].


Biodiversity policy evaluation can help preserve biodiversity by increasing the effectiveness of conservation policies. Improved policy evaluation can be measured by changes in • • • • •

the number of government agency personnel that have received training in conservation policy evaluation [no.]; government support for establishing or maintaining regional centers of excellence in conservation policy evaluation [$]; the existence of requirements for evaluation of conservation policies by implementing government agencies [0/1]; the rigor of requirements for evaluation of conservation policies by implementing government agencies [qualitative]; and the number of rigorous evaluations of conservation policies conducted [no.].

4.6 Policies omitted from lines of action Of the policies discussed in Section 3, four do not appear in the lines of action: ecocertification, ecotourism, bioprospecting, and CSR programs. In each case, the evidence that the policy is generally effective in conserving biodiversity and/or generally does not pose significant threats to biodiversity is thin.1 As noted in the introduction to this chapter, this is not to say that these four policies cannot be effective in conserving biodiversity or avoiding harm – clearly that is not true – just that in our judgment, evidence of general effectiveness or lack of harm is thin. The lack of evidence could be due to a paucity of credible evaluations (a pervasive problem, as discussed in Sections 3.16 and 4.5) and/or to a lack of positive results from credible evaluations. This evidence, or the lack of it, is discussed in detail in Sections 3.9, 3.10, 3.11, and 3.14; we only briefly recapitulate it here. In the case of ecocertification, few empirical studies of its environmental effects have been published, and the

Lines of action


only two rigorous ones fail to find a significant benefit. As for ecotourism, the issue is that, notwithstanding any benefits, it can have significant adverse effects by increasing human presence and infrastructure in already-sensitive areas. Bioprospecting clearly can have significant economic benefits. However, an emerging consensus holds that on a per hectare basis, these benefits are unlikely to provide sufficient incentives for preservation of biodiversity. Finally, an extensive body of research on CSR demonstrates that, on average, it has few additional environmental benefits.

Note 1 Note that the full list of criteria in Table 4.1 does not necessarily apply here; the criteria pertain to broad lines of action, not more specific conservation policies.


Latin America and Caribbean biodiversity actors

Appendix 3 lists major institutional actors in LAC biodiversity conservation in nine categories: 1 2 3 4 5 6 7 8 9

national protected areas agencies; national ministries of the environment; multilateral agencies; bilateral cooperation and other foreign government agencies; foundations; international nongovernmental organizations; research centers; Consultative Group for International Agricultural Research centers; and members of RedLAC (see below).

RedLAC, the Latin American and Caribbean Network of Environmental Funds, is a regional network of national environmental trust funds and other organizations. It seeks to mobilize funds for biodiversity conservation and provide crucial information for conservation funders to evaluate what others have done. With funding from the Gordon and Betty Moore Foundation, RedLAC administers the Ecofunds Database, which provides users with information on environmental investments in the region, including information on individual grants, funders, and implementers.

Appendix 1 Stakeholder interviews

Alpízar, Francisco Director, Environment for Development Center for Central America, Centro Agronómico Tropical de Investigación y Enseñanza (Costa Rica) March 16, 2012 Biggs, Christopher Lead Natural Resource Economist, United Nations Development Programme (Panama) March 30, 2012 Bovarnick, Andrew Lead Natural Resource Economist, United Nations Development Programme (Panama) March 30, 2012 Ferraro, Paul Associate Professor, Georgia State University March 23, 2012 Kaimowitz, David Director of Sustainable Development, Ford Foundation March 30, 2012 Muñoz-Piña, Carlos Director Resource Economics, Forest Trends and Distinguished Visiting Professor, Universidad Iboamericana March 30, 2012 Narain, Urvashi Lead Environmental Economist, the World Bank March 30, 2012 Ramos, Aurelio Senior Conservation Strategist, The Nature Conservancy (TNC) March 23, 2012


Appendix 1

Rodríguez, Carlos Manuel Director, Mexico and Central America, Conservation International (CI) March 23, 2012 Windehoxel, Nestor Program Director MAREA, US Agency for International Development (USAID) March 30, 2012

Appendix 2 Country-level data

This appendix contains tables of country-level data meant to complement subregion-level tables and statistics in the body of the book. It is provided for those interested in specific LAC countries. In cases where the appendix tables are simply country-level versions of subregion-level tables in the body of the book, the numbering of the two tables is the same except that the appendix tables include an “A” prefix (e.g., Table A2.1.1 is a country-level version of Table 2.1.1 in the body of the book. In cases where the appendix tables do not have an analogue in the body of the book, they include the letter “S” for supplement (e.g., Table A2.2-S1).


Appendix 2

Table A2.1-1 Forest area in LAC, by country, type and use, 2010 Country

Forest area (1,000 ha)

Primary forest area (1,000 ha)

Antigua and Barbuda Argentina Bahamas Barbados Belize Bolivia Brazil Chile Colombia Costa Rica Cuba Dominica Ecuador El Salvador Grenada Guatemala Guyana Haiti Honduras Jamaica Mexico Nicaragua Panama Paraguay Peru Dominican Republic Saint Kitts and Nevis Saint Vincent and the Grenadines Saint Lucia Suriname Trinidad and Tobago Uruguay Venezuela Total

10 29,400 515 8 1,393 57,196 519,522 16,231 60,499 2,605 2,870 45 9,865 287 17 3,657 15,205 101 5,192 337 64,802 3,114 3,251 17,582 67,992 1,972 11 27 47 14,758 226 1,744 46,275 946,756

– 1,738 0 0 599 37,164 476,573 4,439 8,543 623 0 27 4,805 5 2 1,619 6,790 0 457 88 34,310 1,179 0 1,850 60,178 – – 0 12 14,001 62 306 –

Note n.s.: not significant, indicating a very small value. Source: FAO 2011.

Appendix 2


Forest area (percentage of land area)

Primary forest area (percentage of forest area)

Production forest (percentage of forest area)

Planted forest (percentage of forest area)

22.7 10.7 51.4 18.6 61.1 52.7 62.4 21.7 54.5 51.0 26.1 60.0 35.6 13.9 50.0 33.7 77.2 3.7 46.4 31.1 33.3 25.7 43.7 44.3 53.1 40.8 42.3 69.2 77.0 94.6 44.1 10.0 52.5 47.2

– 6 0 0 43 65 92 27 14 24 0 60 49 2 14 44 45 0 9 26 53 38 0 11 89 – – 0 24 95 28 18 –

– 5 – 0 0 0 7 46 13 14 31 – 2 24 1 28 97 54 21 2 5 20 14 n.s. 37 – 0 – 0 27 34 64 49

– 5 0 1 n.s.* n.s. 1 15 1 9 17 n.s. 2 5 1 5 0 28 0 2 5 2 2 n.s. 1 – – n.s. 3 n.s. 56 –


Appendix 2

Table A2.1-2 Trends in forest area, by country, 1990–2010 Country

Annual change in total forest cover 1990–2000

Antigua and Barbuda Argentina Bahamas Barbados Belize Bolivia Brazil Chile Colombia Costa Rica Cuba Dominica Ecuador El Salvador Grenada Guatemala Guyana Haiti Honduras Jamaica Mexico Nicaragua Panama Paraguay Peru Dominican Republic Saint Kitts and Nevis Saint Vincent and the Grenadines Saint Lucia Suriname Trinidad and Tobago Uruguay Venezuela Source: FAO 2011.


1,000 ha/yr


1,000 ha/yr


n.s. –293 0 0 –10 –270 –2,890 57 –101 –19 38 n.s. –198 –5 0 –54 0 –1 –174 n.s. –354 –70 –42 –179 –94 0 0 n.s.

–0.3 –0.88 0 0 –0.63 –0.44 –0.51 0.37 –0.16 –0.76 1.7 –0.55 –1.53 –1.26 0 –1.2 0 –0.62 –2.38 –0.11 –0.52 –1.67 –1.18 –0.88 –0.14 0 0 0.27

n.s. –252 0 0 –10 –271 –3,090 42 –101 23 52 n.s. –198 –5 0 –54 0 –1 –120 n.s. –235 –70 –12 –179 –94 0 0 n.s.

–0.4 –0.81 0 0 –0.65 –0.46 –0.57 0.26 –0.16 0.95 2.06 –0.57 –1.73 –1.43 0 –1.32 0 –0.74 –1.95 –0.1 –0.35 –1.91 –0.35 –0.94 –0.14 0 0 0.23

n.s. 0 –1 49 –288

0.64 0 –0.3 4.38 –0.57

n.s. 0 –1 22 –288

0.13 0 –0.31 1.48 –0.59

Appendix 2


Annual change in primary forest cover 2005–2010




1,000 ha/yr Percentage 1,000 ha/yr Percentage 1,000 ha/yr Percentage 1,000 ha/yr Percentage 0 –240 0 0 –10 –308 –2,194 38 –101 23 35 n.s. –198 –4 0 –56 0 –1 –120 n.s. –155 –70 –12 –179 –150 0 0 n.s.

0 –0.8 0 0 –0.68 –0.53 –0.42 0.23 –0.17 0.9 1.25 –0.59 –1.89 –1.47 0 –1.47 0 –0.77 –2.16 –0.12 –0.24 –2.11 –0.36 –0.99 –0.22 0 0 0.3

– 0 0 0 0 –176 –2,812 –10 –14 0 0 n.s. – 0 0 –27 – 0 – n.s. –402 – 0 0 –72 – – –

0 –4 –1 45 –288

0 –0.02 –0.32 2.79 –0.61

0 –7 0 1 –

0 –0.44 –0.54 –0.21 –0.16 0 – –0.3 – 0 0 –1.2 – – – –0.07 –1.07 – – 0 –0.12 – – –

– 0 0 0 0 –176 –2 734 –10 –14 0 0 n.s. 12 0 0 –27 0 0 – n.s. –188 – 0 0 –225 – – 0

0 –0.05 0 0.31 –

0 –9 0 1 –

0 – –

0 –0.46 –0.55 –0.21 –0.16 0 – –0.31 0.26 0 0 –1.32 0 – – –0.07 –0.53 – – 0 –0.36 – – –

– 0 0 0 0 –200 –2,336 –10 –14 0 0 n.s. 12 0 0 –68 0 0 0 n.s. –44 –27 0 0 –177 – – 0

0 –0.53 –0.48 –0.22 –0.17 0 – –0.31 0.26 0 0 –3.72 0 – 0 –0.07 –0.13 –2.16 – 0 –0.29 – – –

0 –0.06 0 0.33 –

n.s. –18 0 1 –

2.03 –0.13 0 0.26 –

0 – –

– 0 – –


Appendix 2

Table A2.1-5 Area devoted to agricultural crops and pasture, by country and year (thousands of ha)* Country

Cropland 1990

Antigua and 9 Barbuda Argentina 27,420 Bahamas 10 Barbados 17 Belize 77 Bolivia 2,255 Brazil 57,408 Chile 3,049 Colombia 5,000 Costa Rica 510 Cuba 3,841 Dominica 16 Ecuador 2,925 El Salvador 810 Grenada 12 Guatemala 1,785 Guyana 502 Haiti 1,100 Honduras 1,820 Jamaica 219 Mexico 26,300 Nicaragua 1,495 Panama 654 Paraguay 2,199 Peru 3,920 Dominican 1,350 Republic Saint Kitts and 10 Nevis 10 Saint Vincent and the Grenadines Saint Lucia 18 Suriname 68 Trinidad and 71 Tobago Uruguay 1,305 Venezuela 3,610 Total 149,795





Percentage 1990 growth, 1990–2008 9


28,900 33,000 20 11 11 10 17 17 0 99 102 32 3,168 3,819 69 65,200 68,500 19 2,110 1,722 –44 4,545 3,461 –31 490 500 –2 4,054 3,970 3 19 21 31 2,979 2,500 –15 900 915 13 11 11 –8 1,965 2,268 27 478 445 –11 1,220 1,300 18 1,427 1,428 –22 250 235 7 27,400 27,500 5 2,151 2,130 42 695 695 6 3,110 4,300 96 4,285 4,440 13 1,318 1,300 –4







99,970 99,870 99,850 2 2 2 2 2 2 49 50 50 33,200 33,831 33,000 184,200 196,206 196,000 12,850 13,000 14,015 40,083 40,314 39,153 1,795 1,350 1,300 2,900 2,500 2,630 2 2 2 4,921 5,087 4,945 600 600 637 1 1 1 2,500 2,500 1,950 1,230 1,230 1,230 497 490 490 1,500 1,508 1756 257 229 229 77,500 78,400 75,000 2,530 2,943 3,016 1,470 1,535 1,535 14,960 17,215 16,100 17,916 16,900 17,000 1,200 1,197 1,200

0 0 0 2 –1 6 9 –2 –28 –9 0 0 6 0 –22 0 –1 17 –11 –3 19 4 8 –5 0


4 –59






8 –20





14 67 60

10 –44 56 –18 47 –34

2 20 6

2 21 7

1 19 7

–50 –6 17

13,520 13,543 13,191 18,250 18,000 18,000 533,941 548,543 542,318

–2 –1 2

1,415 1,673 3,395 3,350 161,778 169,747

28 –7 13

Note *Area devoted to agricultural crops equal to arable land area plus permanent cropland area. Source: ECLAC 2011.

Percentage growth, 1990-2008

Appendix 2


Table A2.1-6 Market value of agricultural output, by country and year (millions of 2005 $)* Country




Percentage growth, 1990–2008

Antigua and Barbuda Argentina Bahamas Barbados Belize Bolivia Brazil Chile Colombia Costa Rica Cuba Dominica Ecuador El Salvador Grenada Guatemala Guyana Haiti Honduras Jamaica Mexico Nicaragua Panama Paraguay Peru Dominican Republic Saint Kitts and Nevis Saint Vincent and the Grenadines Sainta Lucia Suriname Trinidad and Tobago Uruguay Venezuela Total

10 10,480 126 74 51 707 25,699 2,227 8,334 985 3,184 54 1,303 1,397 42 2,198 143 – 828 623 21,159 453 579 963 2,785 1,533 5 40 81 85 104 979 3,779 91,011

11 13,819 149 70 94 947 35,198 3,541 9,751 1467 2,087 48 1,938 1,574 38 2,902 322 – 1,063 662 24,604 709 805 1,211 4,561 1,953 5 31 50 83 124 1,216 4,695 115,728

19 17,522 138 57 109 1,199 50,463 5,417 12,000 1,869 2,077 44 2,852 1,980 30 3,647 294 – 1,408 605 30,507 893 1,142 2,043 6,240 2,448 6 33 37 92 91 1,544 6,023 152,830

98 67 10 –23 113 70 96 143 44 90 –35 –18 119 42 –29 66 106 – 70 –3 44 97 97 112 124 60 33 –18 –54 8 –13 58 59 68

Note *Includes fishing and forestry. Source: ECLAC 2011.


Appendix 2

Table A2.1-7 LAC soybean and sugar cane production, by country and year (thousands of tons) Country

Soybeans 1990


Sugar cane 2007

Percentage 1990 growth, 1990–2007

Antigua and – – – – Barbuda Argentina 10,700 20,136 47,483 344 Bahamas – – – – Barbados – – – – Belize 0.146 0.495 0.377 158 Bolivia 233 1,197 1,596 586 Brazil 19,898 32,735 57,857 191 Chile – – – – Colombia 232 38 51 –78 Costa Rica 0 – – – Cuba – – – – Dominica – – – – Ecuador 167 94 61 –63 El Salvador 2 3 3 25 Grenada Guatemala 42 30 35 –17 Guyana – – – – Haiti – – – – Honduras – 3 2 – Jamaica Mexico 575 102 88 –85 Nicaragua 12 6 2 –82 Panama – 0 0 – Paraguay 1,795 2,980 5,856 226 Peru 3 3 3 5 Dominican – – – – Republic Saint Kitts and – – – – Nevis Saint Vincent and – – – – the Grenadines Saint Lucia – – – – Suriname 0.033 0.025 0.011 –67 Trinidad and – – – – Tobago Uruguay 37 7 780 2,008 Venezuela 4 4 60 1,485 Total 33,699 57,339 113,877 238 Source: ECLAC 2011.



Percentage growth, 1990–2007 –

15,700 18,400 19,200 22 60 63 6 –90 606 546 410 –32 1,089 1,107 1,200 10 3,880 3,602 6,419 65 262,674 327,705 549,707 109 – – – – 27,791 33,400 32,000 15 2,630 3,800 3,561 35 81,800 36,400 11,900 –85 5 4 5 7 5,721 5,402 8,360 46 2,957 5,140 4,956 68 7 7 7 11 9,603 16,552 25,437 165 2,700 2,710 3,099 15 1,500 801 1,000 –33 2,898 3,974 5,958 106 2,491 2,025 1,968 –21 39,919 44,100 52,089 30 2,392 3,524 4,481 87 1,298 1,789 1,798 39 2,449 2,245 3,400 39 6,700 7,535 8,229 23 6,512 4,511 4,824 –26 170


105 –38



20 –17

– 65 1,478

– 120 1,373

– – 120 85 475 –68

683 150 293 –57 6,619 8,832 9,691 46 492,419 536,026 760,719 54

Appendix 2


Table A2.2-S1 Population within 100 km of the coastline, by country Country

Total population

Population within 100 km

Percentage within 100 km

Aruba Anguilla Netherland Antilles Antigua and Barbuda Bahamas Belize Bermuda Barbados Costa Rica Cuba Cayman Islands Commonwealth of Dominica Dominican Republic Guadeloupe Grenada Guatemala Honduras Haiti Jamaica Saint Kitts and Nevis Saint Lucia Mexico Montserrat Martinique Nicaragua Panama Puerto Rico El Salvador Saint Pierre and Miquelon Turks and Caicos Islands Trinidad and Tobago Saint Vincent British Virgin Islands US Virgin Islands Argentina Bolivia Brazil Chile Colombia Ecuador Falkland Islands French Guiana Guyana Peru Paraguay Suriname Uruguay Venezuela

100,572 11,410 209,645 64,848 304,225 229,856 62,960 267,498 4,027,855 11,199,089 38,229 70,559 8,392,521 433,453 93,502 11,376,955 6,437,953 8,122,562 2,576,082 38,473 147,783 98,868,768 3,749 383,385 5,069,790 2,850,683 3,914,679 6,258,172 6,852 16,699 1,294,352 113,279 23,649 120,881 37,136,032 8,332,114 170,360,272 15,203,581 42,092,560 12,648,820 2,316 164,582 760,166 25,656,386 5,438,244 418,090 3,338,939 24,177,932

100,572 11,410 209,645 64,848 304,225 226,875 62,960 267,498 4,020,296 11,199,090 38,229 70,559 8,392,521 433,453 93,502 7,363,292 4,052,839 8,122,562 2,576,082 38,473 147,783 28,547,270 3,749 383,385 3,400,626 2,850,683 3,914,679 6,236,023 6,852 16,699 1,294,352 113,279 23,649 120,881 15,872,910 0 80,793,940 11,974,450 10,817,170 6,357,693 2,316 136,445 566,602 14,688,470 0 392,416 2,499,404 14,719,680

100 100 100 100 100 99 100 100 100 100 100 100 100 100 100 65 63 100 100 100 100 29 100 100 67 100 100 100 100 100 100 100 100 100 43 0 47 79 26 50 100 83 75 57 0 94 75 61

Source: CIESIN 2007.


Appendix 2

Table A2.2-S2 Current mangrove area and loss rates, by country Country

Remaining area (km2)a

Loss rate (%/yr)b

Loss rate (ha/yr)b

Antigua and Barbuda Aruba Bahamas Barbados Belize Brazil Cayman Islands Colombia Costa Rica Cuba Dominican Republic Ecuador El Salvador French Guiana French Polynesia Grenada Guadeloupe Guatemala Guyana Haiti Honduras Jamaica Martinique Mexico Netherlands Antilles Nicaragua Northern Saint-Martin Panama Peru Puerto Rico Saint Kitts and Nevis Saint Lucia Saint Martin Saint Vincent and the Grenadines Suriname Trinidad and Tobago Turks and Caicos Islands UK Virgin Islands US Virgin Islands Venezuela Totale

9.3 0.7 811 0.4 569 10,630 76 2,147 390 4,286 180 1,300 337 857 0 2.1 20.7 353 224 147 670 96 10.8 7,302 3.3 739 0.1 1,538 34.5 83 0.5 1.4 0.1 0.5 745 64 172 0.8 1.9 3,360 37,163 km2

3.60 0.00 0.20 9.20 0.20 0.10 0.30 0.80 1.80 0.00 2.90 0.50 1.50 0.00 0.00 1.30 0.10 0.00 0.20 0.60 3.80 0.70 0.40 1.20 0.60 1.30 0.00 0.70 1.70 0.00 0.90 0.00 0.00 0.10 0.00 0.20 0.00 0.70 5.10 0.60 1.0c (0.57d)

33 0 162 4 114 1,063 23 1,718 702 0 522 650 506 0 0 3 2 0 45 88 2,546 67 4 8,762 2 961 0 1,077 59 0 0 0 0 0 0 13 0 1 10 2016 21,151 ha

Notes a Estimated by Siikamäki et al. (2012a, b) using data from Giri et al. (2010). b Estimated by Siikamäki et al. (2012a, b) using data from FAO (2007). c On average, by country. d Overall in the region. e World total area of mangroves is about 139,000 km2 (Giri et al. 2010).

Appendix 2 Table A2.2-S3 Landings from coastal fisheries, by country, 2008 (tons) Country Antigua and Barbuda Argentina Bahamas Barbados Belize Bolivia Brazil Chile Colombia Costa Rica Cuba Dominica Ecuador El Salvador Grenada Guatemala Guyana Haiti Honduras Jamaica Mexico Nicaragua Panama Paraguay Peru Dominican Republic Saint Kitts and Nevis Saint Vincent and the Grenadines Saint Lucia Suriname Trinidad and Tobago Uruguay Venezuela Total Source: ECLAC 2011.

Landings 3,521 995,083 9,236 3,551 4,621 6,797 775,000 3,939,377 135,002 21,750 27,902 694 434,239 48,000 2,384 22,826 42,168 10,010 12,904 13,175 1,594,338 29,810 222,508 20,000 7,376,686 15,424 450 3,828 1,713 23,811 13,833 110,691 295,364 16,216,696



Appendix 2

Table A3.1-1 Terrestrial and marine area protected, by country and year (percentage terrestrial area or territorial waters
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