.

UNITED STATES 

ENVIRONMENTAL PROTECTION AGENCY 

A FRAMEWORK FOR THE 

ECONOMIC ASSESSMENT OF 

ECOLOGICAL BENEFITS 

February 1, 2002 

Acknowledgements 
The preparation of A Framework for Ecological Benefits Assessment (henceforth Framework) was 

managed under the direction of the Social Sciences Discussion Workgroup, a working group under the 

Science Policy Council. Initial work on the Framework began in 1996 and concluded in 2001. 

The principle managers of the Framework workgroup were Betsy Southerland and Mary Ellen Weber. 

The workgroup consists of staff economists and ecologists from program and regional offices across the 

Agency. The workgroup was originally chaired by John D. Harris and then by Lynne Blake-Hedges. 

Lynne Blake-Hedges was responsible for final preparation of text with valuable assistance from Randy 

Bruins, Matt Heberling, and Anne Sergeant. Contributors to the development of the content and focus of 

the document include Ghulam Ali, Erik Beck, Ed Bender, Richard Healy, Rich Iovanna, Virginia Kibler, 

Robert Lee II, Bill O’Neil, Neil Patel, Rosalina Rodriguez, Elliot Rosenberg, Christine Ruf, and Will 

Wheeler. 

In addition to EPA staff, a number of contractors provided administrative support and developed key 

materials used as technical assistance in preparation of the Framework. They include Carolyn Russell, 

Kevin Blake, and Margaret McVey of ICF, Incorporated. 

Reviews of document drafts were completed by Erik Beck, Ed Bender, Robert Perciasepe, Will Wheeler, 

Joan P. Baker, Jeri-Anne Garl, John Miller, Dan Phalen and Shari Stevens of EPA; as well as Ed 

Whitelaw, University of Oregon; Paul Courant, University of Michigan; Ernie Niemi, ECONorthwest; 

Jeffrey Lazo, Penn State University; James Shortle, and Penn State University. A final peer review was 

conducted in 1998. Critical review and comment provided by Paul Jacobson, Langhei Ecology LLC and 

John Hopkins University, Barbara Kanninen, University of Minnesota, Greg Poe, Cornell, and Jonathan 

Rubin, University of Maine. 
A FRAMEWORK FOR THE ECONOMIC

ASSESSMENT OF ECOLOGICAL BENEFITS

Prepared for 

Ecological Benefit Assessment Workgroup 

Social Sciences Discussion Group 

Science Policy Council 

U.S. Environmental Protection Agency 

February 1, 2002 
.
TABLE OF CONTENTS

Acknowledgements ........................................................... iv

1.0 Introduction ............................................................ 1

1.1 Background ...................................................... 1

1.2 Purpose and Scope ................................................. 1

1.3 Organization of this Document ....................................... 4

2.0 A Proposed Framework ................................................... 5

2.1 Overview ........................................................ 5

2.2 Planning ......................................................... 7

2.3 Problem Formulation – Initial Coordination

.................................................... 9

2.3.1 Problem Formulation ......................................... 9

2.3.2 Initial Coordination ......................................... 10

2.4 Linking Endpoints During Problem Formulation

............................................................... 13

2.4.1 Conceptual Model of a Cascade of Ecological Effects ................ 13

2.4.2 Identifying Preliminary Economic Endpoints

......................................................... 15

2.4.3 Identifying and Defining Linkages

......................................................... 16

2.5 Problem Formulation – Prioritizing Endpoints and Selecting Valuation Techniques

............................................................... 19

2.5.1 Prioritization Criteria ........................................ 19

2.5.2 Monetized, Quantitative, and Qualitative Assessments

......................................................... 21

2.6 Problem Formulation – Ensuring Analytical and Data Compatibility in the

Analysis Plans 

............................................................... 23

2.6.1 Establishing the Baseline and Alternative Scenarios

......................................................... 23

2.6.2 Measuring and Modeling Linkages

......................................................... 24

2.6.3 Matching Spatial Scales

......................................................... 25

2.6.4 Matching Temporal Scales ................................... 26

2.6.5 Data Limitations and Uncertainty

......................................................... 26

2.7 Conducting the Assessments ........................................ 27

2.7.1 Ecological Risk/Benefit Assessment ............................ 27

2.7.2 Qualitative and Quantitative Economic Assessment ................ 28

2.8 Characterizing and Presenting Results ................................ 29

2.9 Concluding Remarks .............................................. 30

Page i 
References and Further Reading ........................................... 31

3.0 Important Principles of Ecology and Ecological Assessment .................... 34

3.1 Defining Ecosystem and Other Levels of Ecological Organization .......... 34

3.1.1 Definitions ................................................ 34

3.1.2 Levels of Biological Organization .............................. 35

3.1.3 Interactions Within Ecosystems ................................ 35

3.2 Understanding Ecosystem Structure and Function ....................... 39

3.3 Valued Ecological Entities .......................................... 43

3.3.1 Definitions ................................................ 43

3.3.2 Identifying Valued Ecological Entities .......................... 44

3.3.3 Neglected Benefits .......................................... 46

3.4 Types of Ecological Assessments .................................... 50

3.4.1 Assessment Models ......................................... 50

3.4.2 Standardized Approaches to Ecological Assessments ............... 53

4.0 Ecological Risk/Benefit Assessment ....................................... 59

4.1 Overview of EPA’s Guidelines for Ecological Risk Assessment ............ 59

4.2 Phase I: Problem Formulation ...................................... 61

4.2.1 Conceptual Model .......................................... 63

4.2.2 Assessment Endpoints ....................................... 66

4.3 Phase II: Analysis Phase ........................................... 71

4.4 Phase III: Risk/Benefit Characterization .............................. 82

5.0 Background Theory on Valuing Changes to Ecological Resources ................ 86

5.1 Welfare Economics and the Value of an Ecological Change ............... 86

5.2 Measuring the Benefits of Improvements to Ecological Resources – The Concept

of Willingness-To-Pay ............................................. 87

5.3 How Economic Benefits of Improvements to Ecological Resources are Realized

............................................................... 88

5.4 Estimating Willingness-to-Pay ...................................... 89

6.0 Economic Assessment of Ecological Benefits ................................. 91

6.1 Components of an Economic Assessment of Ecological Benefits ........... 91

6.1.1 Identify and Prioritize Economic Benefit Endpoints ................ 92

6.1.2 Describe and Quantify Changes to the Economic Benefit Endpoints ... 93

6.1.3 Estimate the Value of the Changes ............................. 93

6.1.4 Summarize and Present the Results ............................. 94

6.2 Identifying the Service Flows and Other Values Provided 

by an Ecological Resource .................................... 96

6.2.1 Direct, Market Uses ......................................... 99

6.2.2 Direct Non-Market Uses .................................... 101

6.2.3 Indirect, Non-Market Uses .................................. 103

6.2.4 Non-Market, Non-Use Values ................................ 105

6.3 Approaches to Measuring Resource Values ........................... 107

6.3.1 Market Price and Supply/Demand Relationships ................. 112

6.3.2 Market-Based Valuation Approaches .......................... 115

Page ii 
6.3.3 Travel Cost Methodologies .................................. 121

6.3.4 Random Utility Model ...................................... 125

6.3.5 Hedonic Price and Hedonic Wage Methodologies ................ 128

6.3.6 Contingent Valuation ....................................... 134

6.3.7 Combining Contingent Valuation with Other Approaches: Contingent

Activity ................................................. 143

6.3.8 Conjoint Analysis and Contingent Ranking ..................... 145

6.3.9 Benefits Transfer .......................................... 150

7.0 Issues Affecting the Economic Valuation of Ecological Benefits ................. 155

7.1 Uncertainty and Variability ........................................ 155

7.2 Discounting .................................................... 156

7.3 Distributional and Equity Analyses .................................. 156

8.0 References ........................................................... 158

8.1 Ecological References and Further Reading ........................... 158

8.2 Economic References and Further Reading ............................ 168

Page iii 
Acknowledgements 

The preparation of A Framework for Ecological Benefits Assessment (henceforth Framework) was 

managed under the direction of the Social Sciences Discussion Workgroup, a working group under the 

Science Policy Council. Initial work on the Framework began in 1996 and concluded in 2001. 

The principle managers of the Framework workgroup were Betsy Southerland and Mary Ellen Weber. 

The workgroup consists of staff economists and ecologists from program and regional offices across the 

Agency. The workgroup was originally chaired by John D. Harris and then by Lynne Blake-Hedges. 

Lynne Blake-Hedges was responsible for final preparation of text with valuable assistance from Randy 

Bruins, Matt Heberling, and Anne Sergeant. Contributors to the development of the content and focus of 

the document include Ghulam Ali, Erik Beck, Ed Bender, Richard Healy, Rich Iovanna, Virginia Kibler, 

Robert Lee II, Bill O’Neil, Neil Patel, Rosalina Rodriguez, Elliot Rosenberg, Christine Ruf, and Will 

Wheeler. 

In addition to EPA staff, a number of contractors provided administrative support and developed key 

materials used as technical assistance in preparation of the Framework. They include Carolyn Russell, 

Kevin Blake, and Margaret McVey of ICF, Incorporated. 

Reviews of document drafts were completed by Erik Beck, Ed Bender, Robert Perciasepe, Will Wheeler, 

Joan P. Baker, Jeri-Anne Garl, John Miller, Dan Phalen and Shari Stevens of EPA; as well as Ed 

Whitelaw, University of Oregon; Paul Courant, University of Michigan; Ernie Niemi, ECONorthwest; 

Jeffrey Lazo, Penn State University; James Shortle, and Penn State University. A final peer review was 

conducted in 1998. Critical review and comment provided by Paul Jacobson, Langhei Ecology LLC and 

John Hopkins University, Barbara Kanninen, University of Minnesota, Greg Poe, Cornell, and Jonathan 

Rubin, University of Maine. 

Page iv 
1.0 INTRODUCTION 

1.1 BACKGROUND 

The Social Sciences Discussion Group (SSDG), convened under the auspices of the EPA 

Science Policy Council, was initiated to address issues related to the conduct of economic and 

other social science analyses at EPA. One of their efforts focused on improving the Agency's 

ability to conduct economic benefit analyses for regulatory cost-benefit or relative benefit 

assessments. The SSDG identified a need to "improve the Agency's ability to quantify, and, 

where possible, monetize ecological benefits, including quality of life." A workgroup 

representing all major EPA programs and environmental media was established to meet that 

charge. 

The workgroup began by surveying EPA offices for completed or ongoing analyses of ecological 

benefits to determine the current state of the practice within EPA. During this exercise, the 

workgroup identified the need for a common approach to analyzing ecological benefits and a 

better understanding of both the scientific and economic techniques used in these analyses. 

1.2 PURPOSE AND SCOPE 

This document represents a joint effort of ecologists and economists. This document is intended 

to address the two needs identified above by (1) proposing a common framework for the 

economic analysis of ecological benefits and (2) discussing the elements of ecological risk 

assessment and economic benefit analysis. In addition, this document is intended to: 

C Promote greater coordination between ecologists and economists working on such 

efforts; 

C Provide an understanding of the approaches and techniques currently in use; 

C Suggest additional sources of information; and 

C Provide a starting point for individuals who need to assign economic values to changes in 

ecosystems that have or might result from human activities. 

This document is intended to provide general information to EPA staff and others who are 

interested in the concepts and techniques used to assess and quantify ecological effects of an 

environmental decision and to monetize ecological impacts and benefits. An important aspect of 

the document is an introduction of a framework for collaboration between economists and 

ecologists. The framework presented is not intended as Agency guidance and should not be 

considered to be promoting any particular benefits methodology. 

The document is not designed to be either a “cookbook” or a “how to” manual — it does not 

provide a step-by-step guidance on the application of specific techniques. Because this 

document is a framework for estimating the economic value of ecological benefits, it also does 

Page 1 
not address other possible effects of an action or other perspectives. Specifically, this document 

does not

 discuss non-ecological effects, such as human health impacts or socio-economic 

effects (e.g., employment, local revenue, growth). 

The Framework presented in this document represents one part of a larger process of 

environmental decisionmaking at EPA, as illustrated by the white box with a double-line border 

in Exhibit 1. The goal of the Framework is to provide a structure for conducting benefits 

assessments for the purpose of informing risk management decisions and to meet risk 

management objectives. It is most applicable for determining, as part of a benefit cost analysis, 

the ecological benefits of policies or regulatory actions commonly undertaken by governmental 

agencies such as the EPA. Other types of analyses that might be conducted to inform a decision, 

such as human risk assessment, environmental justice assessments, and other types of economic 

assessments, are beyond the scope of this document, but may be included in the decisionmaking 

process. Discussion of other types of economic analyses can be found in EPA’s (2000) 

Guidelines for Preparing Economic Analyses. Other activities included in environmental 

decisionmaking, such as monitoring and program evaluation, also are not addressed in this 

Framework. 

For an economic analysis of ecological benefits of a decision to be conducted, an assessment of 

the ecological changes that might result from that decision is needed. In essence, an ecological 

risk assessment is required, where both beneficial changes as well as potentially adverse changes 

(the usual implication of the word “risk” in many risk assessment contexts) in ecological 

endpoints, are evaluated. We have therefore chosen to develop a framework for the economic 

evaluation of ecological benefits around the framework in EPA’s (1998) Ecological Risk 

Assessment Guidelines, which consists of planning, problem formulation, analysis of stresses and 

effects, and risk characterization. 

The Framework proposed in this document identifies the major points in the ecological and 

economic assessment processes where coordination between the two assessments is needed. 

This Framework is intended to provide a starting point for approaching such analyses; it does not 

prescribe a particular method of research or interaction. An important aspect of the Framework 

is the recognition that ecological risk assessment and economic benefits assessment are distinct 

disciplines. It focuses on the phases of the respective assessment processes during which 

communication and coordination between ecologists and economists are needed to ensure an 

adequate benefits assessment. While it has been assembled based on the experience and 

judgment of EPA environmental economists and ecological risk assessors, this Framework per se 

has not undergone testing – such as through a program of case studies. 

Efforts to manage ecological resources using a place-based (or community-based) approach can 

differ from a government agency-based approach and may require additional tools not described 

in this document. Place-based environmental management situations may be characterized by 

multiple parties; varied or competing objectives; weak or decentralized authorities; and a broad 

range of potential actions which analyses must seek to narrow and define. Economic approaches 

involving scenario simulation, multiple agents, or multiple objectives, as well as conventional 

valuation approaches, may be useful. An EPA, Office of Research and Development effort to 

Page 2 
integrate place-based ecological risk assessment and economics, which explores the utility of a 

range of economic tools, is underway (U.S. EPA, 2000b). In addition, an EPA and Science 

Page 3
Exhibit 1 

Environmental Decisionmaking Process 

Advisory Board workshop exploring the potential contribution of non-economic approaches to 

ecological valuation, including approaches from psychology, anthropology and decision science, 

occurred in May 2001 and will be followed by a report. 

Page 4 
1.3 ORGANIZATION OF THIS DOCUMENT 

The remainder of this document is organized in seven sections. 

Chapter 2: The Framework: Provides an overview of the Framework and a description of 

each of its components, emphasizing the points of coordination between the 

ecological risk assessment team and the economic assessment team. 

Chapter 3: Important Principles of Ecology and Ecological Assessments: Defines 

ecosystems and biological levels of organization. Describes the interactive nature 

of ecosystems and cascading effects. Provides an overview of prospective and 

retrospective assessments. 

Chapter 4: The Ecological Risk Assessment Process: Provides an overview of EPA’s 

Framework and guidelines for ecological risk assessment, emphasizing where 

coordination with the economists is needed to ensure an adequate economic 

assessment of possible ecological changes. 

Chapter 5: Background Theory on Valuing Changes to Ecological Resources: Provides 

an introduction to how economists define the value of ecological resources and 

the theoretical basis for estimating changes in these values to measure economic 

benefits. 

Chapter 6: Economic Assessment of Ecological Benefits: Provides an overview of the 

economic assessment of ecological benefits, including detailed information on the 

types of benefits that might be identified and the techniques available for valuing 

changes. 

Chapter 7: Issues: Discusses some additional issues relevant to the economic analysis of 

ecological benefits, including uncertainty, discounting, aggregation, and equity. 

Chapter 8: References: Provides a complete listing of the materials used to develop the 

Framework as well as suggested readings for additional information. 

References and Further Reading 

U.S. EPA. 1998. Guidelines for Ecological Risk Assessment. Risk Assessment Forum. 

EPA/630/R-95/002F. 

U.S. EPA. 2000a. Guidelines for Preparing Economic Analyses. Office of the Administrator. 

EPA 240-R-00-003. September. 

U.S. EPA. 2000b. Research Plan for Integrating Ecological Risk Assessment and Economics in 

Place-Based Decision Making. NCEA-C-0633. National Center for Environmental Assessment, 

Office of Research and Development, Washington, DC. 

Page 5 
2.0 A PROPOSED FRAMEWORK

This section describes a proposed framework for the economic assessment of ecological benefits. 

It describes the relationship between ecological and economic analyses, identifying many of the 

points of interdisciplinary coordination between the ecological risk assessment team and the 

economic analysts. This chapter provides an overview of this Framework (Section 2.1) and a 

discussion of the key phases and elements of the process, including the planning phase (Section 

2.2), several elements of the problem formulation phase (Sections 2.3 through 2.6), conducting 

the analyses (Section 2.7), and characterizing and presenting results (Section 2.8). 

2.1 OVERVIEW 

The proposed Framework describes a process by which the ecological risk assessors and 

economic analysts conduct and coordinate their assessments as recommended by recent EPA 

guidance. (See U.S. EPA, 1997; 2001.) Interdisciplinary coordination promotes: 

C Development of better information for risk managers; 

C Greater utility of ecological assessments for economic assessment of ecological benefits; 

C Greater relevance of economic assessments to ecological resource issues; and 

C Timely and streamlined collection of necessary data. 

Exhibit 2 illustrates this Framework within the context of the larger environmental 

decisionmaking process. While this Framework follows the general phases outlined in EPA’s 

(1998) Guidelines for Ecological Risk Assessment, it focuses on the coordination between the 

ecologists and economists during those phases. EPA’s ecological risk assessment process can 

readily be adapted to an ecological benefits assessment and consists of a problem formulation 

phase, analysis phase, and risk characterization phase (U.S. EPA, 1998). The steps of the 

economic benefits analysis, as identified in EPA’s (2000) Guidelines for Preparing Economic 

Analyses, are matched with the ecological risk/benefit assessment process. The economic 

benefits assessment consists of identifying the potentially affected benefit endpoints, quantifying 

the significant changes to these endpoints, and estimating the economic value of those changes. 

As a precursor to conducting an economic assessment of ecological benefits, a planning step 

occurs where risk managers, stakeholders, risk assessors, economists, and other parties each 

share their perspectives on the problem to help guide planning of goals, scope, and resources for 

the assessment. Planning for and beginning an economic benefit analysis simultaneously with 

the ecological benefit assessment allows for the coordination called for by the Framework and 

can greatly improve the economic assessment of ecological benefits. The majority of the 

interdisciplinary coordination between the ecological risk assessors and the economic analysts 

occurs during the planning and problem formulation phases and again at the end, in the 

Page 6 
Exhibit 2 

A Framework for Economic Assessments of Ecological Benefits 

A

Exposure 

Planning 

Dialogue among Risk Managers, Stakeholders, Risk Assessors, and Others 

Problem 

Formulation 

Analyses 

Characterize 

Results 

Present

 Results 

Economic Assessment of Ecological Benefits 

Ecological Assessment Economic Assessment 

Conceptual Model 

Endpoints - Linkages 

Analytic Compatibility 

Analysis Plan Analysis Plan 

Problem Formulation Problem Formulation 

Other Analyses 

s needed, Acquire 

D

ata, Iterate 

Process, M

o

nitor Results 

Economic Analyses 

Monetary 

Quantitative 

Qualitative 

Ecological 

Analyses 

 Effects 

Risk Characterization 

Present Results 

Communicating Results to Risk Manager/Environmental Decisionmaker 

Risk Management and 

Communicating Results to Interested Parties 

presentation of the results. The planning and problem formulation steps of the economic 

assessment of ecological benefits include the following: 

• During the planning phase of the overall assessment, the planning team agrees on the 

scope of the proposed action, the risk management objectives, the management 

alternatives and policy options that will be explored by the assessments, the types of 

studies and activities that will occur as part of each assessment, and the basic information 

Page 7 
required to support each of those studies and activities (Section 2.2). This document only 

discusses planning as related to the economic assessment of ecological benefits. 

• At the beginning of problem formulation, ecologists and economists discuss the 

information agreed to during the planning phase as it relates to their analyses and begin 

the design of their respective conceptual models (Section 2.3). 

• After identifying the direct and possible indirect ecological changes that might result 

from options under assessment, economists and ecologists work together to link 

ecological changes to economic endpoints (Section 2.4). 

• Based on the resources available for the assessment and other appropriate criteria, 

ecologists and economists prioritize the assessment endpoints to identify a subset for 

quantitative analysis (Section 2.5). 

• As ecologists and economists develop their assessment plans, they confer on several 

issues to ensure that the ecological assessment and the economic benefit analysis are 

analytically compatible before finalizing their assessment plans at the end of problem 

formulation (Section 2.6). 

• The ecological risk assessment team conducts its assessment. Using results from the 

ecological risk assessment, economists complete their assessment of the economic 

benefits of the ecological changes (Section 2.7). 

• Both groups characterize and present the results of their assessments (Section 2.8). The 

ecological risk assessment team might also present results for some questions unrelated 

to the economic analysis (e.g., which species might serve best as indicators for future 

monitoring of ecological changes). 

Because portions of the ecological risk/benefit assessment and economic analysis of the 

ecological changes will occur simultaneously, along with any other studies included in the 

overall assessment, certain parts of the process outlined in this Framework might be repeated 

until the information is sufficiently precise to be of use in decisionmaking. 

2.2 PLANNING 

Environmental decisionmaking begins with a planning step. The goal of planning is to identify 

the context of the environmental decision, the risk management objectives, the options under 

assessment, the individuals involved, the types of analyses that are needed, what resources are 

available, and to resolve other questions concerning scope and process. Typically, an ecological 

risk management objective has an entity, an attribute, and a desired state or direction of 

preference. The risk management objective helps to focus the assessments on risks that are 

susceptible to management. The decisionmakers or risk managers meet with staff who can 

represent the disciplines that might be required, including toxicologists, sociologists, economists, 

ecologists, risk assessors, engineers, as well as potentially interested or affected parties. The 

assessment teams might include representatives of Federal, State, local, and tribal governments, 

Page 8 
commercial, industrial, and private organizations, leaders of constituency groups, and other 

sectors of the public. 

Several types of questions may be addressed during the planning phase: 

• What kind of decision is involved (e.g., siting a facility; regulation development)? 

• Who is involved in decisionmaking (i.e., who are the risk managers)? 

• Who are the stakeholders (those affected and those who are needed to help define the 

value of ecological benefits)? 

• What analyses will be required (e.g., socioeconomic, human health, ecological)? 

Products of planning dialogues can include several types of information: 

• The type of management goals at issue (e.g., benefits exceed costs, restore striped bass 

population in the Potomac to its 1940 level); 

• The exact nature and timing of the decisions to be made and who will make them; 

• The analyses to be performed and which analyses must be coordinated; 

• Agreement on scope, complexity, and focus of the assessments, including the expected 

output; and 

• The timeline and technical and financial support available to conduct the assessments. 

Depending on the context, decision, management goals, and other attributes of the decision at 

hand, one or more of several different types of analyses might be called for, including human 

health risk assessment, ecological risk assessment, economic assessment of health risks, 

economic assessment of ecological changes, environmental equity or justice, and others (Exhibit 

1). This Framework addresses only the economic assessment of ecological benefits, and thus 

focuses on a subset of the possible analyses that might be conducted during the environmental 

decisionmaking process. 

During planning, the economists, ecologists, toxicologists, and staff representing other involved 

disciplines can contribute to a definition of the scope of the problem and alternatives to be 

considered. For example, economic analysts might provide preliminary information from market 

research efforts that indicate what chemicals/products are being manufactured and used and in 

what amounts. Producers and consumers and chemical uses can be identified. Risk assessors 

might use this information to determine where (geographically) exposures might take place, the 

magnitude and frequency of exposure (how chemicals/products are used), and exposed 

populations (both human and ecological). 

Economic analysts might also provide insights on economically feasible substitutes that might be 

used if a chemical or product is no longer available, projecting potential market shares and 

production and use volumes. Socioeconomic information on the conditions of affected 

communities and/or populations can provide insight as to what the concerns and potential risk 

management alternatives are for a particular community. Economic factors, in addition to others 

such as political and cultural factors, also come into play when identifying feasible risk 

management alternatives. 

Page 9 
Ecologists can provide information on likely direct and indirect ecological impacts, sensitive 

ecosystems or receptors, and likely spatial and temporal scales over which ecological changes 

might be expressed. With information on environmental persistence and bioaccumulation 

potential for chemicals, ecologists can evaluate whether food-chain analyses might be important 

to the assessment. They also can provide initial judgments on likely magnitude, severity, and 

reversibility of anticipated effects. 

By the completion of the planning step, agreements have been made about several aspects of the 

assessment: (1) the management goals, (2) the range of management options that the ecological 

and economic analyses are intended to support, (3) characterization of the decisions to be made 

given the management goals, (4) the focus, scope, and complexity of the assessments, and (5) 

resources available to conduct the assessment. At a general level, there must be agreement on 

the spatial and temporal scale of the assessments. At this point, the groups that will need to 

coordinate with each other also are identified. 

2.3 PROBLEM FORMULATION – INITIAL COORDINATION 

2.3.1 Problem Formulation 

Once planning is complete, the risk and economic assessments can begin. In its (1998) 

Guidelines for Ecological Risk Assessment, EPA describes problem formulation as the first 

phase of an ecological risk assessment. The same principles described in those Guidelines can 

apply to problem formulation phase for an economic assessment. Problem formulation in this 

Framework includes integrating available information, selecting ecological and economic 

assessment endpoints, developing of a conceptual model linking the proposed actions to 

ecological and economic endpoints, and developing analysis plans for the ecological and 

economic benefit assessments (Exhibit 3). 

Page 10 
Exhibit 3 

Problem Formulation for Economic Assessment of Ecological Benefits 

ic 

Identify Endpoints and Linkages Between Them 

Ecological 

Prioritize Endpoints and Select Valuation Techniques 

Develop Analysis Plans, Ensuring Analytical Compatibility 

Develop Conceptual Models Econom

Ecological Analysis Plan Economic Analysis Plan 

Integrate Available Information 

This document focuses on four parts of the problem formulation phase: (1) initial coordination in 

defining the problem; (2) developing the conceptual models that identify linkages between 

ecological and economic endpoints; (3) prioritizing endpoints for quantitative and qualitative 

assessment, and (4) ensuring analytic compatibility between the ecological risk and economic 

assessment plans. Exhibit 4 provides a detailed illustration of the problem formulation phase. 

Questions Asked During Problem Formulation 

! What laws already protect what entities? 

! What are the policy considerations (law, corporate stewardship, societal concerns, intergenerational 

equity)? 

! What is the nature of the problem? 

! What is the context of the assessment? 

! What is the likely scale of the assessment in time and space? 

! What is the starting information like? 

! Develop a conceptual model of the problem, context, scale, then: 

-What are the most obvious assessment endpoints (i.e., relevance to management goals)? 

-What additional endpoints are ecologically and socially relevant (e.g., laws can be good indicators of 

societal values) or important to stakeholders? 

2.3.2 Initial Coordination 

Page 11 
During the preliminary dialogues among the risk managers, risk assessors, economic analysts, 

and stakeholders in the planning stage, the ecological assessors and economic analysts will have 

begun to formulate ideas about the tasks ahead of them. Fundamental to beginning problem 

formulation is a clear description of the environmental decision at hand, the options under 

evaluation, the types of actions, and the initial changes those actions would cause in the abiotic 

environment (e.g., land use, water flow, chemical concentrations). The ecologists will begin to 

formulate an understanding of the proposed options and the types of ecosystem structures and 

services that might be affected by those options; the economists will develop some ideas as to 

the most obvious economic benefits that might accrue. The discussions initiated during the 

planning stage between ecologists and economists to define the scope of the overall assessment 

will continue throughout the problem formulation process. The sharing of information and ideas 

can be particularly helpful in developing comprehensive conceptual models. 

For both the ecological and economic assessments, problem formulation begins by integrating 

the information from the initial discussions of the problem, including the context of the 

assessment, its goals and constraints, the decisions to be made, and which stakeholders are 

involved. Separate ecological and economic assessment teams are assembled to include the 

required areas of expertise and also possibly to include stakeholder representatives. Careful 

consideration about who will participate and how they will participate is best done up front. 

Page 12 
Exhibit 4 

Problem Formulation 

i i i

l l l 

1.

/ ( )

) 

(circles) 

i i i

; ; state 

; ;

i i

Identify Linkages between Ecolog cal Change and Econom c Benefit Endpo nts 

Estab ish Preliminary Conceptua Mode

 From planning: context, goals, constraints, 

decisions, decisionmakers, stakeholders 

2. Integrate available information 

3. Develop conceptual model linking 

action source rectangle to direct ecological 

effects (ovals and indirect ecological effects 

Establish Preliminary Economic Model 

1. From planning: context, goals, constraints, 

decisions, decisionmakers, stakeholders 

2. Integrate available information 

3. Identify likely economic benefit endpoints in 

different categories (squares) and develop 

conceptual model 

Direct Market Direct NonUse Market Use 

Indirect Non­ Non-Use Value 

Market Use 

Pr or tize Ecological Endpoints Prioritize Econom c Endpoints 

Agree to Assessment Endpoints for Analysis 

Develop Preliminary Analysis Plan 

For assessment endpoints, develop measures; 

models characteristics of ecosystems

variables spatial and temporal boundaries 

Design Approach for Assessment 

Specify technique(s) for valuing changes to 

selected benefit endpoints; select preliminary time 

horizon discounting; etc. 

Ensure Analyt cal Compatibil ty and Finalize Analysis Plans 

1. Adjust time horizon and time steps of both assessments to match as possible 

2. Adjust spatial extent and resolution of assessment to match as possible 

3. Specify treatment and analyses of uncertainty 

Ecologists Economists 

Page 13
2.4 LINKING ENDPOINTS DURING What Is An Endpoint? PROBLEM FORMULATION 

Endpoints differ by discipline. Ecological

After the initial coordination, each assessment endpoints are explicit descriptions 

assessment team begins to develop an of the actual environmental attribute that is 

explicit conceptual model of their part of expected to change in response to an action.

the analysis. The ecologists begin by Ecological assessment endpoints are 

tracing the consequences of the proposed operationally defined by an ecological entity 

actions from the sources through the initial and its attributes. Changes to ecological 

changes produced in the physical and assessment endpoints are estimated from 

chemical characteristics of the analyses of both direct and indirect effects of

environment, direct effects on ecological the action and in the context of a benefits 

entities, and then the cascade of secondary assessment, are used to estimate changes in the 

ecological effects that might follow. economic benefit endpoints. Economic benefit 

endpoints are the goods or services provided or

The economic benefit assessment is based supported by the ecological resource, directly

on the premise that actions affecting the or indirectly, that have economic value (see

state of an ecological resource, measured Section 5.3) to society, such as recreational

in terms of changes to the ecological fishing.. Changes in the economic benefit 

assessment endpoints, will result in endpoints are used to assess the economic 

changes to the goods and services provided value of the action under study.

by that resource (i.e., changes to the 

economic benefit endpoints). Because of 

this connection, economists need to work 

with ecologists and other scientists in 

determining what economic benefit endpoints are likely to be affected and estimating the 

magnitude of those effects. By working with economists to define the economic benefit 

endpoints, ecologists can help ensure that ecologically significant but less obvious or less direct 

effects are not overlooked by the economic benefit analysis. Furthermore, as ecologists gain a 

better understanding of the objectives and process of the economic benefit analysis, they might 

be able to provide information and data that are better suited to the needs of the economist. 

2.4.1 Conceptual Model of a Cascade of Ecological Effects

The ecologist outlines ecological changes that might result from one or more decisions and 

actions. Exhibit 5 illustrates a simple, preliminary conceptual model that might be drawn up to 

depict possible ecological benefits of improving local septic systems, one of many possible risk 

management actions. The diagram in Exhibit 5 is a substantial oversimplification of a 

conceptual model for purposes of illustration only. For additional examples and explanation of 

the development of conceptual models for ecological risk assessments, see Chapter 4. 

The conceptual model traces the sequence of changes from the initial direct effect of reduced 

nutrient loading to surface waters to the consequences of that effect, here described as reduced 

eutrophication (i.e., nutrient enrichment) of local waters, which in turn would lead to improved 

Page 14 
aquatic habitat for fish and Exhibit 5 shellfish populations, and that, in Diagram Linking Action to Cascade of Ecological Effects turn, might support larger 

populations of breeding and Improved Local Septic Systems migratory shorebirds. The 

reduced eutrophication also 

would be expected to improve Reduced Nutrient Loading the condition and areal extent 

of wetland vegetation and the 

ability of that vegetation to Reduced Eutrophication in Local Waters filter sediments, nutrients, and 

contaminants out of the water 

before the water reaches rivers 

and lakes. Improved Aquatic Improved Wetland 

Habitat Function and Structure 

The cascade of indirect effects 

and interactions among the 

affected entities, which is Increased Macroinvertebrate Abundance characteristic of ecosystems, Improved Water can be difficult for economists Filtration 

and non-ecologists to envision. Increased As the ecological risk Fish/Shellfish 

assessment team develops their Populations conceptual model, they should 

meet frequently with the 

economic analysts to explain 

the ecological relationships Increased Increased Improved and interactions represented by Migratory Bird Shore Bird Water the conceptual model. Visitation Rate Population Quality Communicating with 

economists during the process 

of developing the ecological conceptual model, starting with the most simple preliminary model, 

will greatly improve economists’ understanding of the ecological assessment and help 

economists to define appropriate economic benefit endpoints. 

One difference between the ecological conceptual models for risk assessments designed to 

identify environmental concentrations of concern or to set cleanup levels and those conducted to 

support economic analyses deserves note here. A conceptual model used to design an ecological 

risk assessment to establish environmental concentrations of concern or cleanup levels at a 

specific site often focuses on the most sensitive and exposed (i.e., vulnerable) receptors or 

processes in the ecosystem at the site. The goal is to identify contaminant concentrations in 

environmental media that are unlikely to cause adverse ecological effects. In Exhibit 5, the most 

vulnerable endpoint might be shellfish populations. To document an environmental 

concentration of concern, for example, an ecological risk assessment might be able to stop there, 

and consider only increased shellfish populations as a goal. A conceptual model that will be 

linked to an economic assessment of ecological benefits should attempt to identify all of the 

direct and indirect effects of an environmental decision. The additional ecological effects 

Page 15 
depicted in Exhibit 5 are added to allow the economic analysis to capture a more complete 

spectrum of ecological benefits. 

When developing a conceptual model to support an economic assessment of ecological benefits, 

the ecologists should also consider the possibility for adverse effects. What is beneficial for one 

species can be detrimental to another, as has sometimes occurred when land is managed for the 

benefit of a single species (e.g., game species such as deer). 

2.4.2 Identifying Preliminary Economic Endpoints 

EPA primarily relies on an effect-by-effect approach for estimating the benefits of a policy 

option (U.S. EPA, 2000a). This approach involves identifying the major beneficial effects of an 

action (e.g., various types of improvements to activities or functions of ecological resources), 

assessing the economic value of each of these improvements independently, and summing up the 

individual values to provide an estimate of the total benefits. Identifying the major beneficial 

effects that will be examined in detail in the benefit analysis involves several steps. Before 

economists determine what effects will be examined in the benefits analysis, they first try to 

identify all possible effects. Based on early discussions with ecologists during planning and 

early in the problem formulation stage, economists begin identifying potential economic benefits 

by thinking about the action under study, reviewing analyses of similar actions, and working 

with the ecological assessment team and their preliminary conceptual models to understand what 

ecological changes are expected. 

Economists might identify various types of benefits stemming from changes to ecological 

resources. The economic benefit endpoints are generally viewed as services or uses provided by 

ecological resources. The types of benefit endpoints include direct market uses, direct nonmarket uses, indirect non-market uses, and non-use values. Chapter 5 describes this 

categorization of economic benefit endpoints and provides example services and uses that might 

be considered. This categorization of potential economic benefit endpoints reflects how directly 

each service or use is experienced by an individual and the extent to which an individual can be 

restricted from enjoying the service or use. Characterizing the economic benefit endpoints in 

this way helps economists identify appropriate valuation techniques for each endpoint. 

One method of identifying economic benefit endpoints is to develop a table that links likely 

ecological changes to impacts on human uses and values (e.g., see U.S. EPA, 1995; King, 1997, 

and the U.S. Army Corps of Engineers (USACE) approach described in Cole et al., 1996). For 

example, groundwater discharges contribute to the flow or stock of water in wetlands, streams, 

rivers, and lakes. As a result, a policy that changes the quality or quantity of groundwater might 

affect the services provided or supported by these surface water resources, such as drinking 

water supply and recreational boating, fishing, and hunting. The economist looking at changes 

to groundwater thus might list increased availability of drinking water, increased opportunities 

for river recreation, or improved quality of recreational fishing as potential economic benefit 

endpoints. 

Exhibit 6 provides a simple illustration of how such an approach might be used to develop a 

preliminary list of economic benefit endpoints. Some of the potential economic effects listed in 

Page 16 
Exhibit 6 are not very specific and may need to be refined before the economic value of the 

effect

 can be estimated, but this listing provides a starting point for economists to work with ecologists 

in identifying the economic benefit endpoints that are linked to the policy or action. As 

discussed in the next subsection (2.4.3), the Framework presented here recommends using the 

conceptual model developed for the ecological risk assessment to identify appropriate economic 

endpoints and to facilitate coordination between ecologists and economists. 

Exhibit 6 

Hypothetical List Linking Ecological Changes to Potential Economic Effects 

Ecological Change 

Reduced turbidity of water body 

Increased wetland acreage 

Economic Effects 

Increased commercial and recreational fish harvests 

Reduced water treatment costs 

Improved aesthetic quality of the water 

Reduced costs of storm damage 

Improved recovery after storm-induced combined sewer 

overflows 

Reduced water treatment costs 

Increased commercial and recreational fishery and shellfish 

harvests 

2.4.3 Identifying and Defining Linkages 

With the preliminary work described above, the ecologists and economists can now work 

together to extend the conceptual model developed by the ecologists to include corresponding 

economic benefit endpoints that might be affected. The thoroughness of the economic benefit 

analysis depends on identifying and defining as many of the linkages between changes to 

ecological resource(s) and changes to the economic benefit endpoints as possible. Identifying 

and defining these linkages begins with a qualitative understanding of the relationships and 

interactions that occur within the natural system. As noted earlier, the ecological risk assessment 

team can help the economist understand these relationships by communicating with the 

economist during the development of the conceptual model. The conceptual model should 

clearly identify the direct and indirect effects of an action and form the basis for explaining the 

ecological cascade of effects to the economic analysts. 

Page 17 
The first attempt to link economic benefit 

endpoints to ecological endpoints is likely to 

result in four types of findings: (1) good 

matches between some ecological changes and 

some economic benefit endpoints; (2) a series 

of ecological changes that the economists had 

not considered; (3) a series of potential 

economic valuation endpoints for which there 

are no clear connections to the ecological 

changes represented in the conceptual model; 

and (4) a series of endpoints whose economic 

benefits are ambiguous or unmeasurable. 

Economists can then consider the additional 

ecological changes represented in the 

conceptual diagram and identify appropriate 

economic benefit endpoints to reflect those 

changes. Ecologists can consider whether 

they have overlooked any effects suggested by 

the economic endpoints and whether some of 

the ecological endpoints could be redefined to 

provide a more clear connection to the 

remaining economic endpoints. 

Importance of “Obscure” Ecological Changes 

The process of linking economic benefit endpoints 

to ecological endpoints can be challenging. 

Economic valuation expresses benefits in terms of 

human values. Improvements considered 

important by ecologists (e.g., increased 

biodiversity of a macroinvertebrate stream 

community, see Exhibit 5) might not necessarily 

be appreciated by the public. Therefore, it can be 

helpful to describe the cause-and-effect 

relationship between seemingly unimportant 

ecological changes and changes with obvious 

implications for humans. For example, an 

increase in fishing opportunities can result from 

increased fish populations that occur because the 

macroinvertebrate community is healthier. 

Indirectly, then, the change in the 

macroinvertebrate community is valued by the 

increase in fishing opportunities. 

Example of How this Step Might Work 

It might be reasonable for economists to list 

recreational and commercial fishing as a 

potential economic benefit endpoint for any 

change in water quality. To define linkages 

between ecological and economic endpoints, 

economists work with ecologists to determine 

if a proposed change in water quality will 

actually have any impact on fish populations. 

Once a link is identified, the nature of the 

relationship needs to be defined. For 

example, ecologists and economists might 

discuss which fish species are most sensitive 

to the change (e.g., game fish such as trout), a 

threshold for effects, and the relationship of 

the magnitude of the change to likely 

population size. 

The ecologists and economists continue 

to work together to refine the match 

between ecological and economic 

endpoints. When a new connection (or 

linkage) is identified, the economic 

valuation endpoint is added to the 

expanded conceptual model along with a 

diagrammatic explanation of the 

connection to one or more of the 

ecological changes represented in the 

ecologists’ conceptual model. At this 

point in problem formulation, the goal of 

collaboration is to be all inclusive and to 

extend the conceptual diagram to include 

as many linkages between ecological 

changes and economic benefit endpoints 

as is reasonable. It is an iterative 

process, as the dotted lines looping from 

the joint conceptual diagram back into 

the individual disciplines in Exhibit 4 

indicate. 

The expanded conceptual model 

produced through this collaborative 

Page 18 
process should identify the economic benefit endpoints that are likely to be affected and the 

pathways by which these effects are realized. Exhibit 7 provides an example of how the 

conceptual model presented in Exhibit 5 might be expanded to include economic benefit 

endpoints. For example, stormwater protection is added to the outcomes of improved wetland 

function and structure, which in turn reduces property losses during storms. Other economic use 

endpoints are linked to several ecological effects. Exhibit 7 illustrates the linking of economic 

benefit endpoints to ecological effect endpoints for a very limited set of endpoints. This 

simplified example considers only a single change, reduced nitrogen loading from local septic 

systems, and only some of the potential linkages between the ecological effect of reduced 

eutrophication and changes experienced by some of the economic benefit endpoints. The 

iterative process of refining the list of economic benefit endpoints and their links to ecological 

effects endpoints continues until agreement is reached that the important elements of the problem 

are represented. 

Exhibit 7 

Expanding a Conceptual Model to Include Linkages to 

Specific Economic Benefit Endpoints 

Improved Local Septic Systems 

Reduced Nutrient Loading 

Reduced Eutrophication in Local Waters 

Improved Aquatic Improved Wetland Function 

Habitat and Structure 

Increased Macroinvertebrate 

Abundance 

Improved Storm Improved Water 

 Protection Filtration Increased

Fish/Shellfish

Populations

Increased 

Migratory Bird 

Visitation Rate 

Increased 

Shore Bird 

Population 

Improved Water 

Quality 

Increased Improved Bird Reduced Improved Recreational 

Recreational Watching Property Swimming and Boating 

Fish/Shellfish Landings Opportunities Losses Opportunities 

Page 19 
The economists and ecologists also should consider whether there are any feedback loops 

between ecological and economic endpoints. For example, if improved water quality results in 

increased fish populations and increased recreational fishing, the potential effects of overfishing 

on the same fish populations should be considered. 

For each economic benefit endpoint, economists and ecologists must define the linkages and 

relationships between the ecological changes and the economic endpoints in sufficient detail to 

move forward with prioritization of the endpoints and development of the analysis plans. In 

defining the linkages, the economist must gather sufficient information from the ecological risk 

assessment team to be able to estimate the potential magnitude of the change to each economic 

endpoint and to determine what techniques might be appropriate for estimating the monetary 

value of the change. 

2.5 PROBLEM FORMULATION – PRIORITIZING ENDPOINTS AND SELECTING 

VALUATION TECHNIQUES 

Time and resource constraints generally require that the ecological and economic benefit 

analyses focus on fully explaining and quantifying changes to only a limited number of 

endpoints. This section discusses where coordination between ecologists and economists is 

needed to prioritize ecological and economic endpoints for analysis. 

2.5.1 Prioritization Criteria 

From the ecologist’s perspective, EPA has defined several criteria that can help identify the most 

important and useful endpoints for an ecological risk or benefit assessment (U.S. EPA, 1998): 

• Ecological relevance of an endpoint (e.g., importance to maintaining ecosystem structure 

or function); 

• Susceptibility of the endpoint to the proposed actions. Susceptibility depends on the 

sensitivity of the endpoint to the action (e.g., plants are particularly sensitive to 

herbicides) and on the likelihood of exposure (e.g., is there a pathway by which the stress 

can reach the organisms?); and 

• Relevance to the management goals established during the planning phase. 

EPA’s (1998) Guidelines for Ecological Risk Assessment discuss these criteria, and examples are 

provided in Chapter 4 of this document. Sometimes, the endpoints that best fit the ecological 

selection criteria may not be those best suited for economic valuation. The ecologists might 

include endpoints that the economists do not plan to address (e.g., which ecosystem process 

offers the highest signal-to-noise ratio for purposes of monitoring change after an environmental 

decision is implemented). 

The prioritization of benefit endpoints for the economic benefit analysis requires consideration 

of several factors: 

Page 20 
• Type of information required by decision-makers; 

• Expected magnitude of the change in the economic value of one benefit endpoint relative 

to other endpoints; 

• Anticipated uncertainty associated with the predicted change and value of the change for 

the benefit endpoint relative to other endpoints; 

• Variation in the change to each benefit endpoint under alternative policy scenarios; and 

• Analytical feasibility considerations. 

Economists generally want to estimate the dollar value of those changes that represent the 

greatest economic benefits. However, economists must also consider whether there are likely to 

be significant differences in the change to the benefit endpoint under alternative policy options 

and whether stakeholders or decision-makers will need information on the benefit endpoint even 

if the magnitude of the economic value of the change is relatively small (U.S. EPA, 2000a). If 

the purpose of conducting the benefit analysis is to choose between alternative policy options, 

the ecological changes that experience the greatest variation in economic value across policy 

options might be the most important to address in detail in the economic assessment of benefits. 

Similarly, there may be particular benefit endpoints that are of interest to stakeholders or 

decision-makers that are given priority for detailed consideration in the economic assessment of 

benefits. 

The potential economic value of the change to each economic endpoint will depend on the 

magnitude of the ecological change or changes linked to that economic endpoint. The ecological 

risk assessment team can help economists determine which benefit endpoints are likely to 

experience the largest or most wide-spread changes or vary most significantly across policy 

options. In some cases, an ecological change that is relatively small in magnitude may provide 

large economic benefits. By working with the ecologists to develop the expanded conceptual 

model and to prioritize endpoints, the economists can be sure that those ecological changes are 

included in the conceptual model. Similarly, ecologists can ensure that economists do not 

overlook ecological changes that might appear to be relatively minor but in fact have widespread 

or long-term consequences. Thus, by working together, the ecologists and economists can make 

sure that the “joint” conceptual model encompasses a comprehensive suite of economic benefit 

endpoints. 

What can be measured in the ecological assessment will dictate, in part, what ecological changes 

and economic effects are captured by the economic benefit analysis. Economists need to 

understand the data traditionally collected and developed by an ecological assessment and 

determine how well these data address the data needs of the economic analysis. Ecologists also 

need to better understand the data needs of the economic benefit analysis. Better communication 

between the disciplines during problem formulation allows both the economists and the 

ecologists to identify opportunities to slightly expand the scope of the conceptual model, 

possibly adding additional ecological endpoints or altering slightly the types of information 

developed to provide for significant improvements in the economic benefit analysis. 

Page 21 
Changes to endpoints that are better understood and more certain are given higher ranking than 

changes to endpoints that are less well understood or more variable. Economists want to provide 

a more certain estimate of the benefits of an action to better support policy decisions. However, 

where changes are potentially very large, they need to be considered even though they might be 

highly variable or not well understood. 

The number of benefit endpoints that can be evaluated in detail in the economic benefits analysis 

depends on the type of assessment conducted for those endpoints as well as the time and 

resources available for the economic assessment. Toward the end of this prioritization process, 

the ecological and economic benefit endpoints have been roughly ranked according to ecological 

and economic importance. The ecologists and economists will need to confer and compare their 

rankings of linked endpoints. Where the rankings agree (e.g., an ecological endpoint ranked as 

high priority has an explicit linkage to an economic endpoint that also is ranked as high priority), 

the discussions will be short. Where the rankings disagree (e.g., an ecological endpoint is listed 

as high priority, but the linked economic benefit endpoint is listed as low priority), further 

discussion might help one or the other group change their ranking. Or, the ecologists might 

decide to evaluate a high priority endpoint even though the economists will devote little attention 

to it. Any high priority economic endpoints will need the supporting ecological analyses to be 

conducted. At the end of this step, both groups have identified those endpoints on which they 

will focus their efforts. 

The next section discusses the decision criteria used to determine if a monetary, quantitative, or 

qualitative economic benefit assessment is appropriate and how communication with the 

ecological assessment team supports that decision. 

2.5.2 Monetized, Quantitative, and Qualitative Assessments 

The following issues are considered in determining whether a monetized, quantitative or 

qualitative assessment of the economic benefits associated with each endpoint is appropriate: 

• Need for a dollar value estimate of the ecological benefits associated with the action; 

• Availability of appropriate economic assessment techniques (e.g., techniques for non-use 

values currently not available); 

• Compatibility of available benefit assessment techniques with the data and outputs of the 

ecological assessment; and 

• Availability of relevant ecological and economic data. 

The appropriate type of assessment of the economic value of ecological changes is often 

determined by answering the questions posed during the planning process: "Why is the analysis 

being conducted?" "What are the questions the analysis will address?" and “What are the 

decisions that these analyses will inform?” In some cases, a qualitative or quantitative 

assessment of the economic value of ecological benefits, rather than a monetized assessment, 

may be all that is necessary to support the decisionmaking process. 

Page 22 
If a monetary estimate of economic benefits is needed, economists must determine if they can 

provide a monetized measure of value of the change for each economic endpoint within the time 

and resource constraints of the overall analysis. The ability of economists to provide a 

monetized measure of benefits associated with any particular endpoint depends on the 

applicability of economic valuation techniques to the situation and the availability of the data 

necessary to support the analysis. Most often, the ability of EPA to provide a monetized 

measure of benefits will depend on the applicability of existing value estimates in the literature 

for use in a benefits transfer analysis (see Section 6.3.9 on Benefits Transfer). Once economists 

identify the economic benefit endpoints for which appropriate economic data and techniques are 

available, they can work with ecologists to determine if the ecological assessment can provide 

the information needed on the ecological changes. For example, to apply a value from a 

previous economic study to the current economic benefit assessment (i.e., to use benefits 

transfer) the economist will need a comparable measure of the ecological change as used in the 

original valuation study. 

A thorough economic benefit assessment focuses not just on the effects that can be monetized, 

but on the full scope of effects. Many ecological services are not provided through markets or 

are not readily associated with market transactions. As a result, it may be more difficult or 

impossible to provide a dependable monetized measure of the benefits associated with many 

ecological changes. For those benefits that are not monetized, a qualitative, and when possible 

quantitative, assessment of the economic value of the changes provides a measure of a service’s 

importance and the degree of change, even when a dollar value cannot be assigned to that 

change. For those benefits that are monetized, including a thorough qualitative and quantitative 

discussion of the changes that are valued supports the dollar valued generated by the analysis. 

Again, the ecological assessment team can help economists determine if some type of 

quantitative assessment of the change is possible, or if a qualitative assessment must suffice. 

It might not be possible for the ecological assessment or economic analysis to assess the change 

to some of the ecological endpoints considered key by the ecologists’ selection criteria. For 

example, certain ecological services are too complex and too poorly understood to quantify or 

monetize potential changes (e.g., carbon cycle, nitrogen cycle). Other ecological benefits are 

difficult to characterize and quantify (e.g., species habitat, pollination, microclimate control). 

Principe (1995) termed these “neglected” benefits because they are seldom included in benefit 

assessments. EPA’s resource book on Assessing the Neglected Benefits of Watershed 

Management Practices (U.S. EPA, 2000b), provides further examples of ecological benefits that, 

although important ecologically and economically, are rarely included in benefit assessments 

because they are hard to characterize and quantify. Nonetheless, these services are extremely 

important to our economic and human welfare (Dailey et al., 1997; Dailey, 1997). 

Even though the ecological assessment and economic analysis are not able to estimate a specific 

change, the economic benefit assessment should recognize that the potential impact or 

improvement to the ecological service might have great value to society. In these situations, the 

economic benefit analysis should include a detailed qualitative discussion on the effect of the 

policy or action on the ecological and economic endpoints and, if possible, describe the potential 

economic significance of these changes to society. 

Page 23 
2.6 PROBLEM FORMULATION – ENSURING ANALYTICAL AND DATA 

COMPATIBILITY IN THE ANALYSIS PLANS 

Problem formulation for both the ecologists and the economists culminates in the development 

of the analysis plan for the assessment. Since the ecological assessment is often the main source 

of information for the economist regarding how a specific action or change has affected or is 

likely to affect an ecological resource, it is imperative that the analysis plans are compatible. 

Having compatible analysis plans means having a common understanding of the baseline from 

which effects are measured and the scenarios or policy options to examine, ensuring that the 

outputs of the ecological assessment meet the needs of the economic analysis, defining 

consistent spatial and temporal scales for the analyses, and determining how uncertainty will be 

treated by the analyses. 

Having compatible analysis plans ensures that: 

C Outputs from the ecological assessment are compatible with the needs of the economic 

benefit analysis; 

C Findings of the ecological assessment and economic benefit analysis are analytically 

consistent; and 

C Conclusions of the ecological assessment and economic benefit analysis meet the needs 

of the decisionmakers as defined during the planning stage. 

During the prioritization of ecological and economic endpoints, ecologists and economists will 

have discussed what information is required by the economist and if that information can be 

derived from or developed during the ecological assessment. During the analysis design phase, 

ecologists and economists formalize what information is needed by economists and determine 

how and when that information will be provided. They must also agree on how changes will be 

described or measured (e.g., from what baseline, under what scenarios, at what level of spatial or 

temporal detail) and how any limitations or uncertainties will be represented. The following 

subsections address each of the issues that must be discussed when designing the analysis plan to 

ensure compatibility between the ecological assessment and the economic analysis. 

2.6.1 Establishing the Baseline and Alternative Scenarios 

The baseline from which effects are measured and the specific scenarios or policy options to 

consider must be consistent between the ecological assessment and the economic benefit 

analysis. EPA’s (2000) Guidelines for Preparing Economic Analyses provides detailed guidance 

on specifying the baseline for economic analyses. According to those guidelines, the baseline 

must be appropriate for the question or policy option addressed, identify a particular point in 

time from which point forward the effects of the policy or action are to be assessed, and define 

assumptions about underlying conditions or factors that are unknown or uncertain but will affect 

the conclusions of the assessment (e.g., number of alternative fishing opportunities available). 

Baseline specification might also include determining what alternative assumptions might be 

examined as part of a sensitivity analysis. Because the baseline must be consistent with any 

Page 24 
other analyses conducted as part of the overall study, the parameters for defining a baseline for 

the ecological assessment and economic benefit analysis may be determined during the planning 

stage for the overall study. However, issues relating to baseline specification specifically related 

to assessing changes to ecological resources may arise and must be resolved when developing 

the analysis plan. 

Defining the scenarios to examine includes defining the action or change to be evaluated, the 

area(s) expected to be affected, the time period over which effects will be evaluated, and 

identifying any additional factors or actions (e.g., other regulations) that might affect the 

outcome and determining how they will be accounted for in the ecological assessment and 

economic benefit analysis. Many of these questions will have been addressed during the 

planning stage. However, it may be necessary to reexamine the decisions made during the 

planning stage in light of the joint conceptual model. 

2.6.2 Measuring and Modeling Linkages 

The analysis plans put in writing the assessment design, the analyses that will be conducted, data 

needs, measures, models to be applied, and statistical techniques to use. Both the ecological and 

economic analysis plans should specify what will be measured and how changes in endpoints 

will be expressed. In developing an analysis plan for the economic benefit assessment, 

economists must determine how they will represent the ecological changes in the economic 

analysis. Most economic valuation approaches will require some measure of the ecological 

change associated with the economic endpoint assessed (e.g., an assessment of the value of 

improved swimming opportunities requires a measure of the change in water quality, an 

assessment of improved wildlife viewing opportunities requires information on the estimated 

change in the wildlife population). During the analysis design phase, economists must determine 

if the information required by the economic analysis is specified in the ecological analysis plan. 

If not, the ecologists and economists must confer until agreement is reached on the how 

ecological changes will be characterized at the end of the ecological risk/benefit assessment. 

At this point, an initial screening-level assessment might be planned to (a) assess the likelihood 

of certain linkages between ecological and economic endpoints and (b) determine the sensitivity 

of those relationships. If any of the results are unexpected, the endpoints for assessment might 

be reprioritized. 

To the extent that the conceptual model identified feedback loops between the economic and 

ecological endpoints, the analysis plans must specify how those interactions will be modeled. 

Such interactive models will require more extensive coordination and cooperation between the 

ecologists and economists than models without feedback loops. For example, in assessing the 

economic benefits associated with an increase in fish populations resulting from improved water 

quality, economists may want to account for market adjustments in response to the increased 

supply of fish, namely lower commercial prices and increased consumption. The long-term net 

effect of these reactions may be a slightly smaller increase in fish populations than estimated 

without accounting for this market adjustment. The change in the estimated increase in the fish 

population may have an impact on other ecological resources, such as piscivorus birds, that rely 

on the fish population as their food source. Thus, by accounting for the economic market 

response associated with the commercial fishery, the ecological assessment may provide a better 

Page 25 
estimate of the benefits to the bird populations (which also may be associated with a separate 

economic benefit endpoint - bird watching). 

Until recently, most economic valuation models focused on only a single change, ignoring 

interactions inherent to the natural system and failing to account for interactions between 

multiple economic and ecological endpoints simultaneously. However, with advances in 

ecological and economic models there will be greater opportunities for using models that capture 

the interactions and feedback loops both within the ecological system and between ecological 

and economic endpoints (e.g., incorporate role of human action in ecological model and reflect 

effect of economic changes on human actions). 

In the coordination between the ecological and economic analysis plans, there are two areas that 

require substantial attention: the spatial and temporal scales of the assessments. These must be 

matched, and issues associated with each are described in the next two subsections. 

2.6.3 Matching Spatial Scales 

The benefits of many ecological processes and services “play out” at a much larger spatial scale 

than specific projects under consideration (Limburg, 1999). The proposed actions and 

alternatives often can be delineated geographically and often are limited to small portions of 

watersheds or landscapes. Ecosystems, on the other hand, can be difficult to delineate 

geographically, and the spatial scale of the change in ecological benefits often is much larger 

than the spatial scale of an implemented management practice. Also, because the benefits play 

out at a larger spatial scale, those services are impacted by other projects, land uses, etc., in ways 

that will affect the outcome. All ecosystems lose, gain, or exchange some types of materials and 

energy with neighboring ecosystems through one or more processes. 

The spatial area over which ecological benefits might occur is generally larger or different from 

the spatial area over which the proposed action/alternatives can be delineated because of the 

cascading nature of the ecological effects. One way to overcome this analytic difficulty and to 

ensure a good match between elements of the analyses spatially is to view the natural systems as 

being organized in hierarchies (O’Neill et al., 1986). In this view, the coarser-scale entities (i.e., 

aggregates of finer-scale entities) can be separated into manageable sets of relatively 

homogeneous subgroups (Costanza et al., 1995; Vatn et al., 1999). Watersheds or catchments 

often are the largest unit of analysis for many assessments, with finer scale units being 

comprised of areas with similar slope, soils, vegetation, and microclimate. Once units of 

analysis have been defined in a series of hierarchies, the next analytic task is to maintain the 

necessary finer-scale variations as one moves from the finer- to the more coarse-scale 

entities/aggregations (Vatn et al., 1999). For a discussion of conducting analyses using 

hierarchies based on different spatial scales, see Costanza et al. (1995), Vatn et al. (1999). 

An advantage of developing the conceptual model that describes both the direct and indirect 

ecological effects of an action is that each node in the model (i.e., the endpoint described in a 

box with arrows entering and leaving the box) is likely to be associated with a change in the 

spatial scale of ecological effects. Thus, the ecologists can examine each node and estimate 

whether the spatial scale of effects at that node is likely to be larger than the spatial scale for 

Page 26 
analysis of the preceding node. To increase the utility of the conceptual model, it is helpful to 

provide a description of the geographic scale and location associated with each node in the 

model. 

Defining the spatial area of consideration for the economic analysis is an important step that can 

have significant impacts on the conclusions of the analysis. The spatial area of consideration 

defined by the ecological risk assessment serves as the starting point for defining the spatial 

limits of the economic analysis. Because the economic analysis focuses on human uses 

associated with ecological resources and humans are more mobile than plant and animals, the 

economic analysis might consider a broader spatial area than that defined by the ecological 

assessment. The economic analysis must define an appropriate spatial area from which humans 

may still make use of or otherwise benefit from the services provided by the ecological resource. 

If the economic analysis can monetize benefits for only a portion of the entire area affected, the 

economic assessment’s qualitative discussion should address the entire area affected and 

recognize that the monetary benefits estimate represent only a portion of the benefits expected 

over the entire area. 

2.6.4 Matching Temporal Scales 

As indicated above for the spatial scale, the benefits of many ecological processes and services 

“play out” over a longer time period than a specific project under consideration (Limburg, 1999). 

The activities associated with a management practice might require only weeks or months to 

implement, or might occur at specified points of time each year. The ecosystem responses can 

require years to decades to develop and often are reflected in changes throughout the entire 

annual seasonal cycle. For the purposes of assessing the economic value of the ecological 

changes, effects may need to be assessed over shorter time intervals. A thorough benefits 

assessment needs to consider the role of lagged or future effects and determine how best to 

account for these types of effects. This may include a better characterization of the stream of 

benefits based on scientific information on changes in environmental conditions over time. It 

also may include determining an appropriate discounting scheme for comparing future effects 

against current effects (see Chapter 7 for further discussion on discounting). 

Again, it can be helpful to establish a hierarchy of time steps for the analysis and 

compartmentalize the analysis accordingly (Vatn et al.,1999; Costanza et al., 1995). 

2.6.5 Data Limitations and Uncertainty 

Evaluation of uncertainty should be a theme throughout the analysis phase that follows problem 

formulation. The analysis plans should specify how uncertainty will be addressed. Several 

sources of uncertainty need to be considered. These include human error, natural variability in 

parameters, data gaps, uncertainty about a parameter’s true value, uncertainty introduced by 

models that attempt to predict real-world processes, and other sources. An important distinction 

to maintain throughout the analysis is the difference between natural variability, which can be 

quantified using various statistics, and uncertainties due to lack of information. 

Page 27 
The economic benefits analysis should recognize the uncertainties in the ecological assessment 

process as well as the uncertainties inherent in economic analysis. The level of uncertainty in 

the ecological assessment process as well as the economic valuation process is often substantial. 

As discussed in detail in EPA’s (2000) Guidelines for Preparing Economic Analyses, the issue is 

not to avoid uncertainty but to recognize and account for uncertainty and provide information 

that is useful to decision makers. As noted in the Guidelines, to adequately address uncertainty, 

the analysis should: use the expected or most plausible outcomes; discuss all key assumptions, 

biases and omissions; include sensitivity analyses for key assumptions; and justify the inputs and 

assumptions used based on the results of the sensitivity analyses. (See Chapter 7 for further 

discussion on accounting for variability and uncertainty.) 

2.7 CONDUCTING THE ASSESSMENTS 

Once the analysis plans are complete, the actual analyses can begin. In general, the ecological 

assessment must be conducted first to provide the inputs on predicted changes in ecological 

endpoints for the economic assessment. The ecological exposure and response assessments are 

conducted by the ecological risk assessment team independent of the economists. In other 

words, if problem formulation and planning for the analyses are carefully conducted, there 

should be little need for communication between the economists and the ecologists during the 

ecological analyses. Often, however, unexpected data gaps or unexpected interim modeling 

results might require discussions between the ecologists and economists to resolve such issues. 

2.7.1 Ecological Risk/Benefit Assessment 

During the analysis phase, the ecological risk/benefit assessment team collects the data specified 

in the ecological analysis plan. The team then conducts both an exposure assessment and an 

ecological response assessment. The exposure assessment evaluates the potential sources of 

stress or change, their distribution in the environment, and their overlap with ecological 

receptors. The response analysis attempts to quantify exposure-response relationships and the 

relationship between measures of response and the ecological assessment endpoints. For 

economic assessments of ecological benefits, it is insufficient to identify thresholds for effects, 

as sometimes is done in ecological risk assessments to identify environmental levels of concern 

or cleanup goals for contaminated waste sites. The ecological risk assessment conducted for an 

economic benefits assessment must estimate the type and magnitude of ecological changes to 

allow the economists to predict the economic benefits (i.e., positive economic changes) from an 

environmental decision. 

Because the ecological endpoints can include both direct and indirect effects, the exposure and 

response analyses generally will include two types of assessments. The first is an analysis of the 

relationship between the magnitude and extent of the initial action/stressor and the magnitude 

and extent of direct ecological effects of that stressor. The second is an analysis of the 

relationship between the magnitude and extent of changes in those initial ecological endpoints to 

the magnitude and extent of responses in the endpoints down the cascade of effects depicted in 

the conceptual model. The analysis phase concludes with descriptions of the findings from the 

Page 28 
exposure and response assessments. Additional discussion of this phase of the ecological risk 

assessment is provided in Chapter 4. 

The final phase of an ecological risk assessment is risk characterization (see Exhibit 2). This 

phase integrates the findings from the exposure and response assessments to characterize the 

predicted changes in ecological endpoints. For ecological risk assessments for contaminated 

sites, risk characterization often consists of a simple question. Does the exposure level exceed a 

threshold for effects? For ecological risk assessments conducted to support economic benefit 

analyses, the emphasis of risk characterization needs to be predicting the magnitude and extent 

(both spatial and temporal) of changes in the ecological assessment endpoints. 

Ecological risk assessment conducted to support economic benefit assessments also differ from 

ecological risk assessments for contaminated sites in how uncertainty in the assessment is 

handled. For risk assessments for contaminated sites, data gaps are generally addressed using 

conservative assumptions. Moreover, exposure assessments generally focus on possible highend exposures (e.g., upper 90th percentile). That is because the risk management goal at 

hazardous waste sites often is to be reasonably sure that a site is “clean” (i.e., unlikely to cause 

adverse effects to the assessment endpoints) after site remediation is complete. In other words, 

the assessment is designed to ensure a “reasonable margin of safety”. For an economic benefit 

analysis, on the other hand, uncertainty might be handled in other ways. 

For economic benefit assessments that address large areas (e.g., national, regional, or state 

assessments), best estimates, instead of high-end estimates, often are the most useful for 

characterizing ecological risks. Estimated mean values for the change in an ecological endpoint, 

with some type of confidence interval on those estimates, provides the economists with numbers 

that can be added or multiplied in the economic assessment without compounding conservative 

biases. Lower and upper percentile estimates of the degree of change in an ecological endpoint 

also are useful to the economists. Plausible worst or best case scenarios are generally only 

useful as bounding exercises for the assessment. 

2.7.2 Qualitative and Quantitative Economic Assessment 

Following the effect-by-effect approach to benefit analysis discussed above, economists proceed 

with the qualitative, quantitative, or monetized assessment of changes for each endpoint. 

Economists begin by gathering economic data and developing their models, as called for by their 

analysis plan, in anticipation of the final input from the ecological risk assessment. At this point, 

economists also collect any additional information and data required for their qualitative and 

quantitative assessments. Once the monetized benefits associated with the various economic 

endpoints are estimated, the dollar values are summed together. It is important, however, that 

economists emphasize that the monetized benefits estimate reflects only a portion of the total 

economic benefits expected to accrue from the action. This statement must be supported by 

strong qualitative and quantitative assessments of other benefits not captured by the monetized 

assessment. 

If models or value estimates from other studies are used in the assessment, the analysis must 

describe the source and provide some assessment of the confidence associated with the source. 

Page 29 
For example, if multiple high-quality studies have produced a similar value estimate, the 

economists can have more confidence in using this value estimate in their benefit calculation 

(U.S. EPA, 2000a). 

Chapter 6 discusses in detail the various economic techniques for estimating the economic value 

of changes to different types of economic benefit endpoints. Economists will likely apply more 

than one valuation technique in estimating the total benefits. Care must be taken in designing 

and implementing the economic analysis to avoid double counting of benefits, particularly when 

applying more than one method to estimate the value of changes to related benefit endpoints. 

Additionally, the economic analysis should recognize any negative consequences of the action or 

policy under study that may offset some of the beneficial improvements. 

2.8 CHARACTERIZING AND PRESENTING RESULTS 

For many assessments, ecologists and economists will present their results separately, usually 

with the results of the ecological assessment first. Because characterization of ecological 

risks/benefits provides the input to the economic analysis, it is important that the presentation of 

ecological changes address several factors: the types of ecological changes expected, the 

magnitudes of those changes spatially, temporally, and per unit area (i.e., severity), and the 

certainty associated with those estimates. The presentation of the ecological changes should 

present the conceptual model and assessment endpoints, review and summarize major areas of 

uncertainty and potential bias, discuss the degree of scientific consensus in key areas of 

uncertainty, identify major data gaps, describe any assumptions used to bridge information gaps 

(U.S. EPA, 1998), and indicate how the uncertainty in the results might be magnified through the 

cascade of ecological effects considered. 

In addition to the formal report of the findings for the ecological risk assessment, collaboration 

between the ecological risk assessment and economics teams might help to explain the 

relationships between the ecological endpoints addressed by the ecological assessment and the 

economic endpoints identified in the economic benefit analysis. 

The results from the economic benefit analysis will present the prioritized list of economic 

effects, discuss the criteria used to select the economic endpoints examined in detail by the 

benefit analysis, and discuss how the economic value of the effects was assessed. The monetary 

benefits estimated for some of the changes will be accompanied by the qualitative and 

quantitative assessment of other benefits that were not monetized. If possible, the qualitative 

assessment should discuss the potential magnitude of the economic benefits for any priority 

endpoints that are not accounted for by the quantitative and monetized assessment. The final 

report should also discuss to some degree the other effects identified that were deemed less 

important to the economic analysis. Finally, the results of the economic analysis must disclose 

any source of error in the analysis and the potential impact of such error on the results. The 

presentation of results should identify any possibility of double-counting of benefits, any 

limitations of the analysis, and any potential imprecision and uncertainty associated with the 

benefit estimates. In discussing the potential impact of any source of imprecision or 

uncertainty, economists should discuss whether the analysis is likely to over- or under-estimate 

the economic value of benefits. 

Page 30 
The results presented also may include information and details that are needed for other 

analyses. For example, an equity analysis may require information on the geographic 

distribution of effects, the distribution of ecological effects and economic benefits across 

different ethnic or economic classes of the human population, or the distribution of ecological 

effects and economic benefits over time. 

2.9 CONCLUDING REMARKS 

Economists and ecologists have different views and perspectives that are important to recognize. 

Closer coordination can be encouraged by understanding how the disciplines differ and 

acknowledging these differences. In conclusion, this section identifies some areas in which 

economists and scientists may find they have different approaches or interpretations. 

• Perspective. Economists approach the identification and valuation of changes to 

ecological resources differently than ecologists and other scientists. For example, human 

activities and welfare are the focus of economists while ecologists are concerned with 

complete ecological systems and the interactions between ecological components, which 

may or may not include effects on humans. 

• Terminology. Each discipline has its own terminology, including different units of 

measure. Even common words such as “value,” “benefit,” and “function” have different 

meanings across disciplines. To improve interdisciplinary coordination, care needs to be 

taken to define and use terms consistently. 

• Scale. Part of interdisciplinary coordination is understanding how a change will be 

measured. This requires that ecologists and economists agree on the units of 

measurement and discuss the spatial and temporal boundaries of the analysis. 

• Focus. The ranking of endpoints will likely differ between the ecological and economic 

assessment. Additionally, the approach for assessing changes may differ (e.g., 

economists may measure only the direct change, without addressing system or feedback 

effects). Such differences are partially a consequence of the training associated with each 

discipline but also reflect important differences in the characteristics of the systems 

studied by the respective disciplines. 

• Metrics. Economists focus on the effect of changes to human welfare and typically want 

to standardize effects or welfare changes into dollars to compare effects that may be 

dissimilar. Other metrics may be appropriate for a qualitative and quantitative 

description of ecological and economic effects. 

Page 31 
References and Further Reading 

Ahearn, M.C. 1997. “Why Economists Should Talk to Scientists and What They Should Ask: 

Discussion.” Journal of Agricultural and Applied Economics, July, 29(1): 113-116. 

Bertollo, P. 1998. “Assessing Ecosystem Health in Governed Landscapes: A Framework for 

Developing Core Indicators.” Ecosystem Health 4(1): 33-51. 

Bockstael, N.E., K.E. McConnell, and I.E. Strand. 1989. "Measuring the Benefits of 

Improvements in Water Quality: The Chesapeake Bay." Marine Resource Economics 6(1): 

1-18. 

Cole, R.A., et al. 1996. Linkages Between Environmental Outputs and Human Services, IWR 

Report 96-R-4. Prepared for U.S. Army Corps of Engineers, Evaluation of Environmental 

Investment Research Program. 

Costanza, R (ed.). 1991. Ecological Economics: The Science of Management and 

Sustainability. Columbia University Press, New York. 

Costanza, R., L. Wainger, and N. Bockstael. 1995. Integrated ecological economic systems 

modeling: theoretical issues and practical applications. In: J.W. Milon and J.F. Shogren (eds.), 

Integrating Economic and Ecological Indicators. Practical Methods for Environmental Policy 

Analysis. Praeger, Westport, MA; pp. 45-66. 

Daily, G. C., S. Alexander, P.R. Ehrlich, L. Goulder, J. Lubchenco, P.A. Matson, H.A. Mooney, 

S. Postel, S.H. Shneider, D. Tilman, and G.M. Woodwell. 1997. Ecosystem services: benefits 

supplied to human societies by natural ecosystems. Ecological Society of America, Issues in 

Ecology, Number 2, Spring 1997. 

Daily, G., ed. 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. 

Washington, D.C.: Island Press. 

DeBellevue, E.B., T. Maxwell, R. Costanza, and M. Jacobsen. 1993. “Development of a 

Landscape Model for the Patuxent River Watershed.” Discussion Paper #10, Maryland 

International Institute for Ecological Economics, Solomons, MD. 

Fitz, H.C., R. Costanza, and E. Reyes. 1993. The Everglades Landscape Model (ELM): 

Summary Report of Task 2, Model Development. Report to the South Florida Water 

Management District, Everglades Research Division. 

Fitz, H.C. E.B. DeBellevue, R. Costanza, R. Boumans, T. Maxwell, L. Wainger, and F. Sklar. 

1995. “Development of a General Ecosystem Model (GEM) for a Range of Scales and 

Ecosystems. Ecological Modeling (in press). 

Page 32 
Kaoru, Y., V. K., and J.L. Liu. 1995. “Using Random Utility Models to Estimate the 

Recreational Value of Estuarine Resources.” American Journal of Agricultural Economics, 

February, 77: 141-151. 

King, D.M. 1997. Using Ecosystem Assessment Methods in Natural Resource Damage 

Assessment, Paper #2. Prepared for U.S. Department of Commerce, NOAA, Damage 

Assessment and Restoration Program. 

Limburg, K.E. 1999. Estuaries, ecology, and economic decisions: an example of perceptual 

barriers and challenges to understanding. Ecological Economics 30:185-188. 

Milon, J.W., C. Kiker, and D. Lee. 1997. “Ecosystem Management and the Florida Everglades: 

The Role of Social Scientists.” Journal of Agricultural and Applied Economics, July, 29(1): 99­

107. 

Musser, W.N. 1997. “Why Economists Should Talk to Scientists and What They Should Ask: 

Discussion.” Journal of Agricultural and Applied Economics, July, 29(1): 109-112. 

O’Neill, R.V., D.L. DeAngelis, J.B. Waide, and T.F. Allen. 1986. A Hierarchical Concept of 

Ecosystems. Princeton University Press, Princeton, NJ. 

Principe, P. 1995. “Ecological Benefits Assessment: A Policy-Oriented Alternative to Regional 

Ecological Risk Assessment.” Human and Ecological Risk Assessment 1(4): 423-435. 

Scodari, P. 1992. Wetland Protection Benefits. Draft Report. Prepared for the Office of Policy, 

Planning, and Evaluation, U.S. EPA. Grant No. CR-817553-01. 

U.S. EPA. 1995. A Framework for Measuring the Economic Benefits of Groundwater. Office of 

Water. EPA/230/B-95/003. October. 

U.S. EPA. 1997. A Conceptual Model for the Economic Valuation of Ecosystem 

Damages Resulting from Ozone Exposure. Draft Report. Prepared by Science Applications 

International Corporation, for the Office of Air Quality Planning and Standards, U.S. EPA. 

U.S. EPA. 1997. Guidance on Cumulative Risk Assessment. Washington, DC: Science Policy 

Council. 

U.S. EPA. 1998. Guidelines for Ecological Risk Assessment. Risk Assessment Forum. 

EPA/630/R-95/002F. 

U.S. EPA. 2000a. Guidelines for Preparing Economic Analyses. Office of the Administrator. 

EPA/240/R-00/003. September. 

U.S. EPA. 2000b. Assessing the Neglected Ecological Benefits of Watershed Management 

Practices: A Resource Book. Office of Water. April. 

Page 33 
U.S. EPA. 2001. Risk Characterization Handbook. Memorandum from W. Michael McCabe, 

Deputy Administrator, Office of the Administrator; Washington, DC: Office of the 

Administrator (January 10). 

Vatn, A., L. Bakken, P. Botterweg, and E. Romstad. ECECMOD: an interdisciplinary modeling 

system for analyzing nutrient and soil losses from agriculture. Ecological Economics 30:189-

205. 

Page 34 
3.0 IMPORTANT PRINCIPLES OF ECOLOGY AND 

ECOLOGICAL ASSESSMENT 

This chapter defines some basic terms and concepts used by ecologists and explains how these 

concepts can be applied in conducting an ecological assessment. First, this chapter defines an 

ecosystem and levels of ecological organization and examples of endpoints at each level (Section 

3.1). Next, this chapter describes interactions that occur within ecosystems, including the 

concepts of “food chain,” “food web,” and “energy flow,” competition, predation, and symbiosis 

(Section 3.2). The chapter then examines what attributes of ecosystems and ecosystem entities 

are of value both to society and to sustaining ecosystems themselves (3.3). This chapter 

concludes by describing different standard approaches to ecological assessments and why EPA’s 

ecological risk assessment process is used in the proposed framework for the economic 

assessment of ecological benefits (Section 3.4). 

3.1 DEFINING ECOSYSTEM AND OTHER LEVELS OF ECOLOGICAL 

ORGANIZATION 

3.1.1 Definitions 

Ecosystem Concepts 

According to the Institute of 

Ecosystem Studies “Ecology is the Ecosystems refer to a system formed by the 

scientific study of the processes interaction of a group of organisms and their 

influencing the distribution and environment. An ecosystem may be a pond or the 

abundance of organisms, the entire globe. It can be natural or artificial. All 

interactions among organisms, and ecosystems are composed of components, structure, 

the interactions between organisms and processes (functions). Components are the 

and the transformation and flux of plants, animals, soil, air, and water. Structure refers 

energy and matter.” An ecosystem to spatial and temporal distribution of those 

can be defined in various ways, but components. Processes are the flow of energy and 

one definition that is particularly the cycling of materials and nutrients through space 

useful is “a spatially explicit unit of and time. 

the Earth that includes all of the 

organisms, along with all Ecosystems occur in geographic arrangements. 

components of the abiotic Smaller ecosystems exist within larger ones. The 

environment within its boundaries” scale selected and the boundaries used to define an 

(Likens, 1992). The concept of an ecosystem depend on the problem or question to be 

ecosystem can be applied at any addressed. 

scale ranging, for example, from a 

Source: Appendix A in U.S. Department of the Interior. 1994. small pond to an entire mountain Ecosystem Management in the National Park Service: range. Because ecology is 

concerned not only with organisms 

but with energy flows and material 

cycles on land, in water, and in air, ecology is often defined as the “study of the structure and 

function of nature.” 

Page 35 
3.1.2 Levels of Biological Organization 

There are five levels of biological organization that are conventionally recognized and 

potentially useful in ecological risk/benefit assessments: 

• Individual, 

• Population, 

• Community, 

• Ecosystem, and 

• Landscape. 

Ecological assessments do not address the sub-organismal levels of organization (organ systems 

and cells), nor do they generally address the larger scales of organization of biomes and the 

biosphere. The levels of organization listed above are not defined by the environment. Rather 

they are defined by scientists to facilitate our understanding of relationships within and among 

ecological systems. Thus, these levels can be described as criteria for observation and analysis 

(Allen and Hoekstra, 1992). 

A species is a group of individuals that are able to successfully interbreed. In a species, slight 

biological variations, both genetic and apparent, will exist among individuals. A population is a 

group of organisms of the same species that live in the same place, and have the potential to 

reproduce with one another during their lifetimes. A community is an organized assemblage or 

association of species in a prescribed area or a specific habitat. An ecosystem, defined above 

and described in more detail below, can be viewed as a biotic (i.e., living) community 

functioning within its abiotic (i.e., nonliving) environment. A landscape comprises a group of 

spatially contiguous ecosystems and is usually defined in geographic terms, such as a watershed. 

Ecosystems are often defined in terms of their structural and functional components. Structural 

components are physical elements present in the environment. Examples include soil, nutrients, 

water, and biological entities such as plants, animals, and microorganisms. Functional 

components are processes or interactions that support the structural components, such as nutrient 

cycling and energy flow. It is the pathways of energy flow and cycles of matter that help to 

determine ecosystem boundaries for our purposes of observation. 

3.1.3 Interactions Within Ecosystems 

In an ecosystem, the biological community and the abiotic elements of the environment (e.g., 

water, soil) are bound together by action and reaction, defined by the reciprocal effects of the 

physical environment on an organism and an organism on the physical environment. 

Temperature, moisture, light and other kinds of radiation, texture and chemical composition of 

soil or water, the presence or absence of gases and chemicals, gravity, pressure, and sound can 

all have profound effects on organisms. Examples of interactions between an organism and its 

physical surroundings would be rising river levels forcing muskrats to abandon burrows and 

move to higher ground or the use of sunlight by plants as an energy source. Organisms 

themselves can also affect the physical environment through their activities, thereby indirectly 

affecting other organisms. Examples include beavers damming streams, which changes the 

Page 36 
aquatic community, and earthworms burrowing through and aerating soil, which improves plant 

growth. 

In an ecosystem, interactions also occur among individuals within a population and between 

individuals of different species. For example, the social behavior exhibited by different 

members of a wolf pack is an example of interactions occurring between individuals within a 

population. Predator-prey interactions between wolves and mice are interactions that occur 

between members of different species. 

To understand the influence of the various types of interactions described above on ecosystem 

structure and function, it is important to view more than one level of biological organization. 

Individual-level effects, such as mortality and reduced reproductive success, can have 

population-level effects, such as decreasing or increasing population size and density. On the 

other hand, population-level processes can compensate for individual-level effects. For example, 

in some types of species, increased adult mortality might be compensated for by increased 

survivorship of younger animals to maturity. 

Population Growth. Interactions among individuals within a population and harvesting of 

energy and materials from food allow animals to reproduce and increase in local abundance. 

Population growth rate is a function of birth rates, death rates, time to maturity, and reproductive 

success. Population growth rate also depends on the immigration of individuals from other 

geographic areas into the population and the rate at which individuals emigrate in search of 

better habitat. Reproduction by some species that have just recently invaded a new geographic 

area (e.g., introduced exotic or invasive species) is not hampered by predation or competition for 

resources. Such populations increase in abundance without check (i.e., exhibit densityindependent growth) for some period of time. During that period, the species “intrinsic rate of a 

natural increase” (i.e., the maximal rate at which offspring can be produced) governs the 

population growth rate. Species that mature quickly and produce large numbers of offspring that 

can survive under favorable conditions have a high intrinsic rate of natural increase. Initially, 

when population density is low, population growth might follow an exponentially increasing 

function. Eventually, however, resources available for growth and reproduction will become 

limiting, and birth and death rates will depend on population density. This situation is called 

density dependent population regulation. Such populations are characterized by relatively stable 

population densities (i.e., an equilibrium situation). When population growth rates are density 

dependent, an increase in the rate of loss of juveniles (e.g., harvesting eggs) or adults (e.g., 

hunting) is compensated for (within limits) by increased survival and/or reproductive rates by the 

remaining population. This is the basic principle underlying management of fisheries and game 

populations. 

Life History Strategies. Interactions between a species and its environment over time result in 

the evolution of traits in the species that are adapted to that environment. Some species that have 

evolved in an environment that is densely populated tend to have few offspring in which they 

invest heavily to improve the offspring’s ability to compete with other individuals. Species with 

this type of life history have been called K-selected (Wilson and Bossert, 1971). Facing resource 

limitations, these species have evolved to mature relatively late and attempt to reproduce 

repeatedly (e.g., annual) over their lifespan. Moreover, at each reproductive effort, they produce 

only a few offspring that exhibit a high survivorship. These species also tend to be characterized 

Page 37 
by longer life spans and larger body size. Whales, elephants, and seabirds are good examples of 

K-selected species. At the other end of the spectrum of life history strategies are the r-selected 

species (Wilson and Bossert, 1971). Such species evolve where resources for reproduction and 

growth are not limiting but might exhibit a patchy distribution in the environment. Under these 

circumstances, the best strategy is simply to maximize the number of offspring produced, 

investing little in each of the individual offspring. These species also tend to reproduce one time 

during their life, producing large numbers of offspring that exhibit poor survivorship except 

where they encounter an “empty” patch of habitat. Weeds are classic examples of r-selected 

species. 

Competition. Competitive interactions are those in which two or more species tend to depress 

each other’s population growth rates and abundance (Gotelli, 1998). There are several different 

types of interactions between species that fit this general definition. Exploitation competition 

occurs when species compete with each other for the same resource (e.g., food). For example, 

domestic cattle and bison compete for food (grasses) on open range lands in several areas of the 

mid-western United States. Interference competition occurs when one species interferes with the 

ability of another species to exploit a resource (Gotelli, 1998). An example would be a plant 

species that releases toxic chemicals into the soil, thereby preventing other plants from 

germinating and growing in that soil. Another example of interference competition is introduced 

(i.e., non-native) species of vines that cover other plants, thereby reducing the solar energy 

available to the covered plants, stunting their growth and eventually killing them. Pre-emptive 

competition occurs where species compete with each other over space (Gotelli, 1998). Examples 

include competition for anchorage in the rocky intertidal zone of New England coastal areas by 

barnacles and mussels. 

If environmental conditions were constant over space and time, the population density of those 

species best suited for those conditions would increase at the expense of other species. 

However, environmental conditions vary substantially over time and with geography (e.g., 

altitude, exposure to the sun, soil conditions, rainfall patterns). This environmental variation 

helps to maintain a number of species in competition with each other because the competitive 

edge among species changes as the environmental conditions change. One species might have a 

competitive advantage under some conditions, but be at a competitive disadvantage under other 

conditions. 

Predation. Predation is a direct interaction between two species in which an individual of one 

species consumes an individual of the other species (e.g., as when a hawk captures and 

consumes a rabbit). This interaction results in removal of individuals and biomass from the prey 

species’ population. The relationship between predator and prey often results in oscillations in 

their relative abundances. For example, when a particular prey species prospers under favorable 

environmental conditions, it tends to increase in abundance. The population of predators can 

respond in two ways. In the short-term, the predator can change its behavior (i.e., a functional 

response) and include more of that prey species in its diet. Over the longer-term, the predator 

population abundance can increase as its death rate from starvation decreases and reproductive 

success increases. This increase in predator abundance is self-limiting, however. Increasing 

predation pressure tends to reduce the prey populations, resulting in starvation, reduced 

reproductive success, and a reduction in the abundance of the predator species. This feedback 

system usually results in a reasonably stable equilibrium ratio of predator to prey, although, 

Page 38 
minor to moderate oscillations in that ratio can occur in response to changing environmental 

conditions. 

Some species of animals (e.g., mice, deer) are typically held at densities lower than their food 

resource base could support because of predation. Removal of the predators in such systems 

generally results in increases in the prey population until some other factor becomes limiting 

(e.g., food resources). Where prey are considered pests (e.g., many insect species), removal or 

loss of the predators for any reason can result in an outbreak of the pest species. Such outbreaks 

often cause serious economic losses ingriculture and silviculture. 

Herbivory. Herbivores consume plant materials, usually in a way that does not kill the 

individual plant. Animals that feed on grasses (e.g., bison, cattle) tend to have digestive systems 

adapted for extracting nutrients from very fibrous plant materials and must consume large 

quantities of food to satisfy their metabolic (i.e., caloric) needs. Species that feed on new growth 

(e.g., rabbits), with high protein and low fiber content, can survive on lower total quantities of 

plant material. A different type of herbivory is the consumption of plant seeds, which are rich in 

both protein and lipids. Many species of birds and small rodents specialize in harvesting seeds. 

In general, however, seed eaters (i.e., gramnivores) generally have little impact on the abundance 

of many plant species because of the prolific production of seeds by those plant species. 

Herbivores generally can detoxify a wider array of chemicals than can carnivores (i.e., 

predators). This ability results from an evolutionary arms race between plants and herbivores. 

The plants cannot “escape” from an herbivore; instead, they must develop defenses that work in 

situ. Over evolutionary time, plants have developed a substantial array of defenses, including 

toxic substances stored in their tissues, to discourage herbivores from eating them. The toxic 

substances often cause an herbivore to be acutely ill. Animals that recover from that experience 

will avoid eating that species of plant in the future. In response, however, over evolutionary 

time, herbivores have been developing metabolic pathways to detoxify the toxic chemicals found 

in plants. Carnivores (at least north temperate species), on the other hand, have not needed to 

develop elaborate detoxification pathways because their prey usually cannot afford to use 

chemicals in their body tissues as a defense. 

Pollination. Pollination is a mutually beneficial interaction between a flowering plant and its 

pollinator (e.g., bee, butterfly). The flowering plant offers rewards (e.g., nectar with a high sugar 

content) to attract the pollinators. The flower shape of the plant has evolved so that as the 

pollinator feeds on the reward, it becomes covered in the pollen of that plant. When the 

pollinator moves on to the next individual of that plant species, the pollen can be transferred and 

fertilize the eggs of that individual. Thus, pollinators help plants reproduce sexually, 

maintaining genetic variability among individuals. Many species of plants cannot self-fertilize, 

and thus require their pollinators in order to reproduce at all. 

Symbiosis. Another beneficial interaction occurs in symbiotic relationships, where two 

organisms in close association with each other benefit from the association. Examples include 

the relationship between corals and the algae that grow in their tissues, the relationship between 

an alga and its host fungus to form a lichen (one of the few organisms that can live on bare rocks 

and begin the process of soil formation), between “cleaner” shrimp and the fish that they clean, 

and between acacia ants and the acacia trees they protect in return for nectar and shelter. In 

Page 39 
many symbiotic relationships, death of one member of the pair sooner or later results in illness or 

death of the other. For example, death of the algae living in the tissues of corals, which can be 

recognized by the “bleached” appearance of the coral, often is followed by death of the coral 

organisms themselves. 

In summary, there are many types of interactions within and among species that occur in all 

ecosystems. For this reason, direct impacts (or benefits) of an activity on one species tend to 

produce a cascade of effects through an ecosystem because of the interactions among species in 

the ecosystem. 

3.2 UNDERSTANDING ECOSYSTEM STRUCTURE AND FUNCTION 

Ecosystems may be as large as unbroken tracts of forest and grassland or smaller than a pond. 

The ecosystem is an energy-and-material-processing system, receiving abiotic and biotic inputs. 

The driving force is the energy of the sun. Abiotic inputs include oxygen, carbon dioxide, and 

nutrients. Nutrients become available via weathering of the Earth’s crust and precipitation. 

Biotic inputs include organic materials, such as living organisms and detritus matter (i.e., dead 

and/or decaying organisms). 

The ecosystem itself consists of three components: 

• Producers that derive their energy from the sun (i.e., photosynthetic plants); 

• Consumers and decomposers that use the energy fixed by the producers and eventually 

return nutrients to the ecosystem; and 

• Dead organic material and inorganic substrates that act as short-term nutrient pools and 

support the cycling of nutrients within the ecosystem. 

The most basic functions of the ecosystem are photosynthesis and decomposition. 

Photosynthesis is the process by which green plants utilize the energy of the sun to convert 

carbon dioxide and water into carbohydrates. Through photosynthesis, plants are able to capture 

the sun’s energy and drive the majority of metabolic activities in the living world. Decomposers 

are responsible for the return of nutrients to the ecosystem and a final dissipation of energy to the 

environment. 

Page 40 
Page 41

Exhibit 8

Trophic Level Organization

The “food chain” is a concept that describes the movement of energy and nutrients from one

feeding group of organisms to another in a series that begins with producers and ends with

consumers. The food chain concept specifies a sequence of organisms, each of which feeds on

the preceding group. The trophic structure (“trophic” means “feeding”) of a community is based

on the food chains in the community (see Exhibit 8). A simple food chain might be: oak leaf 4

caterpillar 4 small bird 4 hawk. One useful approach in defining relationships among

organisms is to group organisms based on their trophic levels (i.e., their position in the food

chain). 

The major categories for trophic organization are producers, primary consumers, and secondary

consumers. However, ecosystems are too complex to be characterized by a single, unbranched

food chain. Instead, the transfer of materials and energy from one type of organism to another is

better described as a food web (see Exhibit 9).

The food web for most communities is very complex, including many species and trophic

groups. Several primary consumers may feed on the same plant species. For example, several

insect species might feed on one tree. On the other hand, one species may feed on several

different plants. Also, some species may feed at more than one trophic level. For instance, owls

may eat primary consumers, such as field mice, and also prey on higher level organisms like

snakes. It is more correct, then, to draw relationships between these trophic levels, not as a

simple chain, but as a more elaborate interwoven food web. Complexity of the food creates
Page 42

Exhibit 9

Terrestrial Food Web

Food Web Models

Food webs can be described from various points of view. A sink web is a conceptual model

of a food web that is constructed by tracing trophic links downwards from a single species,

usually a top predator, to the primary producers (Hall and Raffaelli, 1993). This conceptual

model is useful when assessing the potential for a bioaccumulative contaminant in the

environment to cause adverse effects in top predators such as bald eagles and wolves. A

source web is a food web constructed by tracing trophic links upward from a species, using

a basal resource (e.g., a plant species, an insect species) to all those species that feed on it

directly or indirectly (Hall and Raffaelli, 1993). This model can be useful when tracing the

potential effects of pesticides applied to one type of crop. A community web is constructed

by delineating the boundaries of a community, identifying all components of the

community, and determining the feeding links among them. Exhibit 9 is a simplified

example of a community web. Community web models are useful when contaminated

environmental media result in contamination of many different species at the base of a food

web.

opportunities for amelioration of impacts on a particular food chain, but also makes it possible

for there to be indirect effects that are larger than the direct effects.

Two processes occur in an ecosystem through the food web: energy transfer and nutrient cycling. 

Both energy and nutrients are transferred from plants (producers) to herbivores to carnivores

(primary and secondary consumers) and from all preceding levels to the decomposers through

the food web. By tracing the energy transfers and nutrient cycles, the ecologist is able to analyze

the changes in an entire ecosystem. 

Terrestrial and aquatic ecosystems tend to have somewhat different patterns of energy flow and

nutrient cycling. In terrestrial systems, the major energy input is from the sun, while in some

aquatic systems, for example small streams within forests, additional energy inputs come from

the terrestrial environment from plants, insects, and other animals. In forest ecosystems, a

substantial proportion of the organic matter in the system can be in the form of dead and

decomposing organic matter, the leaf litter. Hairston and Hairston (1993) estimate that 95

percent of the net primary production (NPP) by plants in temperate forests reaches the forest

floor, while only about 15 percent of the NPP in a lake reaches the lake bottom. Once on the

bottom, only a fraction of the detritus in a lake is consumed; the rest accumulates annually in the

sediments where it might be permanently “lost.” In terrestrial ecosystems, detritivore-based food

webs tend to be at least as important as herbivore-based food webs, whereas herbivore-based

food webs tend to predominate in aquatic ecosystems (Hairston and Hairston, 1993). 

The length of food chains in terrestrial ecosystems tends to be shorter than the length of food

chains in aquatic ecosystems (Hairston and Hairston, 1993; Oksanen, 1991). Some ecologists

have even proposed that the length of food chains in terrestrial ecosystems typically is two steps,
while in large open water systems (e.g., large lakes), the food chain length typically is three steps 

(Briand and Cohen, 1987; Pimm, 1980, 1982). As the difference in food chain length would 

suggest, bioaccumulation of contaminants in terrestrial ecosystems is less prevalent than 

bioaccumulation of contaminants in aquatic ecosystems. 

Example of Translation and Magnification of the Effects 

of Pollutant Discharges Through the Food Web 

Pollutant discharges can affect not only the health, behavior, and survival of individual 

organisms, but they can also adversely influence the vital interactions and energy flow of the 

food web. This could lead to an adverse change in the structure or function of a population 

or community. When energy and materials flow from one trophic level to the next, 

contaminants in plant or animal tissues consumed are also transferred to the next trophic 

level. If a contaminant is retained in the consumer’s body tissues, its concentration will be 

higher in each succeeding trophic level, because an organism eats many times its own body 

weight during its lifetime. In this way certain pollutants can bioaccumulate as they travel up 

the food web, reaching toxic concentrations in the upper trophic levels of a food web even 

though concentrations in environmental media (e.g., soils, surface water) are relatively low. 

Methylmercury is an example of a bioaccumulative compound. Methylmercury is generally 

present in small amounts in surface waters. It is absorbed directly from the water by aquatic 

organisms, but, more importantly, it bioaccumulates as it is passed up the food web, 

beginning with algae and ultimately passing to fish-eating animals such as gamefish and 

certain mammals, including humans. Mercury concentrations in consumers near the top of 

the food web can reach toxic levels, thousands of times greater than that of the ambient 

Given the discussion above of interactions within ecosystems, it should be clear that defining the 

boundaries of an ecosystem for purposes of assessment can be difficult. Margalef (1968) 

suggested that an investigator might operationally define ecosystem boundaries as locations 

where energy flows are near zero or negligible in comparison with energy flows within the 

system (Suter, 1993). Examples of ecosystem boundaries by this definition would include 

coastlines, forest edges, or watersheds. This definition is consistent with hierarchical 

descriptions of ecosystems, where components within the system are strongly coupled with each 

other and form systems or subsystems that are weakly coupled to other subsystems (Suter, 1993; 

Hoekstra, 1992). A similar definition of ecosystem boundaries might be constructed using 

nutrient cycles instead.3.3 VALUED ECOLOGICAL ENTITIES 

Over the course of many years and activities, EPA has examined the issue of what environmental 

entities should be considered priorities for protection. EPA’s (1997) Priorities for Ecological 

Protection: An Initial List and Discussion Document for EPA states: 

Environmental legislation has a long history of protecting certain groups of animals such as 

fish, shellfish, migratory songbirds and waterfowl, and large mammalian game species. More 

recently, legislation has sought to protect entire ecosystems and to ensure their “integrity” for 

the foreseeable future. 

Page 43 
The following paragraphs describe what is meant by ecological integrity and what characteristics 

of ecosystems are diagnostic of ecological integrity. 

3.3.1 Definitions 

Ecological integrity has been defined by EPA (1994a) as “the interaction of the physical, 

chemical, and biological elements of an ecosystem in a manner that ensures the long-term health 

and sustainability of the ecosystems.” This definition encompasses the concepts of 

sustainability, resiliency, and biodiversity (U.S. EPA, 1997). 

Sustainability is the ability of an ecosystem to support itself over a long time. In the context of 

human use of resources, the concept indicates the ability of an ecosystem to support itself despite 

continued harvest, removal, or other types of losses. For example, the National Marine Fisheries 

Services uses measures of fish reproduction, growth, and recruitment to determine allowable fish 

harvests. 

Resiliency is the ability of an ecosystem to adapt to or to recover from a stress. The stress might 

be natural (e.g., flood, fire, pest outbreak) or anthropogenic (e.g., timber harvest, chemical 

releases, development of land). A system subject to a permanent alteration will not return 

exactly to its original state. Resiliency is a natural attribute of most undisturbed ecosystems, 

reflecting the natural stresses (e.g., drought, temperature extremes, cycles in population 

abundance) to which the ecosystem has adapted over evolutionary time. However, human 

activities often move such stresses beyond the range found naturally and beyond the level of 

resilience developed by the ecosystem during its evolution. 

Biodiversity has been variously defined. One useful definition is that of Norse (1990), who 

describes biodiversity as “the variety of life on all levels of organization, represented by the 

number and relative frequencies of items (genes, organisms, and ecosystems).” Biodiversity is 

one of the keys to an ecosystem’s sustainability and resiliency. Ecosystems that contain more 

species and include higher levels of genetic variation within species often are better able to 

recover from disturbances than other ecosystems. This is because biodiversity tends to reflect 

internal structural and functional redundancies in an ecosystem, such that the loss of some 

individuals or species is compensated for to some extent by other individuals and species (U.S. 

EPA, 1997). 

3.3.2 Identifying Valued Ecological Entities 

Publicly valued ecological functions, services, and entities are evidenced by current laws, by 

private and government actions, and by expressed human values and philosophies. The values 

range from immediate human utility to values that are independent of humans. At one end of the 

spectrum are utilitarian values include the use of natural or manmade resources for direct human 

consumption and in the marketplace. Somewhere in the middle are recreational and aesthetic 

uses and human-derived preservation values. At the other end of the spectrum are those 

associated with moral, religious, and spiritual values (U.S. EPA, 1997). 

Page 44 
For EPA, several statutory provisions direct EPA’s attention to several specific ecological 

concerns. These include ecosystem components, ecosystems, and special places (U.S. EPA, 

1997): 

• Ecosystem components: The Clean Water Act (CWA) specifies fish, shellfish, and 

wildlife as entities for protection.; the secondary National Ambient Air Quality Standards 

(NAAQS) under the Clean Air Act specifies soils, water, crops, vegetation, animals, 

wildlife, etc. as entities for protection.. 

• Ecosystems: The Clean Air Act (CAA) refers to “regionally representative” and 

“critical” ecosystems, and the CWA specifies rivers, lakes, and estuaries. The CAA 

gives the Administrator the authority to assess risks to ecosystems from criteria 

pollutants. 

• Special places: The CWA and CAA identify the Chesapeake Bay, the Great Lakes, and 

Lake Champlain. The CAA also makes special provisions for national parks and 

wilderness areas. The CWA identifies “Outstanding Natural Resource Waters” for 

enhanced protection. 

In its discussion document for Priorities for Ecological Protection, EPA (1997) has proposed 

four criteria for prioritizing ecological entities to be protected (see Chapter 3): mandated 

protection, other societal values, rare or under threat, and ecological significance. 

Mandated protection: Protection for certain types of entities is mandated by law (e.g., 

endangered species are protected by the Endangered Species Act; fish, shellfish, and wildlife by 

the Clean Water Act; and special places such as the Great Lakes by the CWA and CAA). These 

laws codify some of the indirect use and non-use values that humans place on ecological entities. 

Other societal value: As evidenced in its laws, practices, and community projects, society 

values organisms, places, ecosystems, and their structures and functions for commercial, 

recreational, spiritual, or other reasons. Economic assessments regularly address commodities 

that are used in commerce (market) and the recreational (non-market direct-use) values of 

ecosystems. Although techniques to monetize indirect-use and non-use values of ecological 

entities are not yet available, these values can be addressed at least qualitatively in the economic 

assessment. 

Rare or under threat: Species need not be designated as threatened or endangered to be at risk 

of local or regional extinction. Many species of both plants and animals are declining or are 

already so rare that some additional stresses might easily lead to their extinction. An example is 

neotropical migrant songbirds. Although few of these species are federally designated as 

threatened or endangered, populations of most species in this group are declining as their 

habitats in both the northern and southern hemispheres are degraded, fragmented, and lost. 

Even rare species contribute to overall biodiversity and can provide functional redundancies 

within ecosystems, contributing to resilience. Economic valuation of the protection of rare or 

threatened species, communities, and ecosystems can be achieved through the economic 

valuation of biodiversity. 

Page 45 
Ecological significance: Ecological entities that help to sustain ecosystems include plants and 

animals that provide a significant food base, promote nutrient cycling, assist in regenerating 

critical resources, or through competition or predation are “key” to maintaining the balance of 

species in a community. These are often referred to as “keystone” species or functions. If the 

ecological assessment team identifies potential impacts or benefits to keystone species, the team 

should continue the conceptual model to include the important functional and structural 

attributes of the ecosystem or community that might be affected by changes in the abundance or 

presence of a keystone species. 

In response to one of the recommendations of EPA’s (1994b) report Managing Ecological Risk, 

the Agency has begun a process of trying to reach consensus on a list of ecological concerns or 

entities that should be considered in every EPA decision where relevant (U.S. EPA, 1997). 

Exhibit 10, from EPA’s (1997) discussion document Priorities for Ecological Protection, 

provides the proposed list of ecological entities in three categories: (1) animals, plants, and their 

habitats; (2) whole ecosystems; and (3) special places and species. For each entity, the table 

provides examples of attributes of the entities that deserve protection. The combination of an 

ecological entity and an attribute of concern for the entity represents an assessment endpoint. 

Also for each entity, the table provides examples of the management objectives for each specific 

ecological entity. Finally, each ecological entity is evaluated relative to the four criteria listed 

above. 

3.3.3 Neglected Benefits “Neglected” Ecological Benefits

Biotic Resources Of particular importance to economic • Species habitat assessments of ecological benefits is the entity • Biotic productivity “ecosystem functions and services” which have • Species fitness very high ecological significance, but values to • Food chain support

society that often are not recognized. • Biodiversity

• Pest control

Biotic Resources • Pollination

Processes/Infrastructure • A species habitat is defined as the • Microclimate control environment that a given species uses • Geomorphological control over the course of its life history. It • Water supply includes biotic (e.g., assemblages of plant • Energy and nutrient exchange and animal species) and physical (e.g., • Purification of resources

rainfall and temperature range)

components. For animal species, suitable Sources: U.S. EPA, 1993; Principe, 1995 

habitat is essential for both resident and 

transient animal populations. Of 

particular value is habitat for endangered, threatened, and rare species and habitat that is 

vital to important animal species activities (e.g., reproduction, foraging, migration, and 

overwintering). 

• Biotic productivity refers to the total amount of growth among organisms at any level of 

the ecosystem. It includes primary productivity, which accounts for the growth of

Page 46 
autotrophic organisms (primarily plants) that manufacture their own organic materials 

from inorganic sources. Biotic productivity at all trophic levels is essential for energy 

transfer and for maintaining the integrity of natural food webs. 

• Species “fitness” refers to the ability of a species to sustain its populations over the long 

term.1 Attributes that are closely related to species fitness include reproductive success, 

survivorship, and genetic diversity. Genetic diversity within a species is needed to allow 

adaption of the population to changing environmental conditions. 

1 The more traditional use of the word “fitness” is to designate reproductive fitness, which is an 

individual trait, not a species attribute. 

Page 47 
Exhibit 10 

EPA’s 1997 Proposed List of Ecological Entities to be Considered in EPA Decisionmaking 

Category 

Ecological 

entity 

Examples of 

attributes 

Examples of 

objectives 

Criteria 

Mandated 

Societal 

Value 

Rare, under 

threat 

Ecological 

significance 

Animals, plants, and their habitats 

1. Aquatic communities in lakes, streams, and estuaries 

Survival, development, reproduction of aquatic species; habitat extent for key species 

Protect 95 percent of aquatic species, or maintain population of a key species 

Some (CWA) High for fish 

and shellfish 

Some Relatively high 

2. Regional populations of native species and their habitats ­terrestrial and aquatic 

Survival and recruitment; habitat extent 

Maintain viable regional population of native species; maintain or restore habitat for native species 

Some in CAA, CWA 

High for some 

Some High for some 

3. Groups of native or migratory species exposed to severe or acute threat 

Survival without visible damage 

Avoid widespread and recurring or massive die-offs 

Some (e.g. Migratory Bird Treaty Act) 

Usually high Not usually Varies, often 

unknown 

Whole ecosystems 

4. Ecosystem functions and services 

Nutrient recycling, ability to filer pollutants, habitat extent for diversity of species 

Maintain or restore function or service to some standard 

Some general authority 

Not always recognized 

A few Very high 

Page 48 
Exhibit 10 

EPA’s 1997 Proposed List of Ecological Entities to be Considered in EPA Decisionmaking 

Category 

Ecological 

entity 

Examples of 

attributes 

Examples of 

objectives 

Criteria 

Mandated 

Societal 

Value 

Rare, under 

threat 

Ecological 

significance 

Whole ecosystems 

5. Wetlands and stream corridors 

Extent Maintain extent of 

wetland 

Yes High for 

many 

Some High 

Special places and species 

6. Endangered ecosystems (e.g., oldgrowth forests, tall-grass prairies) 

Extent Maintain extent of 

endangered ecosystem types 

Some High for 

some 

All Important for 

biodiversity 

7. Endangered species and their habitats 

Survival, development, reproduction, and recruitment 

Maintain and restore populations 

Yes Potential for 

some 

High Usually low 

8. Other places with high ecological or societal value, as appropriate 

Species diversity; nutrient levels, etc., appropriate to the type of ecosystem; landscape measures 

Restore biodiversity, maintain as oligotrophic lake, maintain extent of certain habitat 

Some (e.g., Great Waters by CAA) 

High for many 

Some High for many 

aThe term objective is used here to refer to a specific objective for an ecological entity.bSpecial places do not necessarily fit the definition of entity in EPA’s (1998) Guidelines for Ecological Risk Assessment.cAn oligotrophic lake is one with low (but adequate) nutrient and high oxygen levels.

Source: U.S. EPA, 1997 

Page 49 
• Food chain support refers to the support of the trophic structure of communities through 

adequate primary productivity and balanced predator-prey relationships that maintain the 

species diversity and abundance of organisms in each trophic level naturally associated 

with an ecosystem. These relationships are essential for energy transfer and for 

maintaining the overall integrity of the food web. 

• Preservation of biodiversity includes maintaining genetic diversity within populations, 

species richness within communities and ecosystems, and ecosystem variety within 

landscapes. Biodiversity is generated and maintained in natural ecosystems, where 

organisms encounter a wide variety of living conditions and chance events that shape 

their evolution in unique ways. Overall, biodiversity provides a reservoir for change, 

enabling life to adapt to changing conditions. 

• Pest control refers to natural pest control, which includes the control of pests by their 

natural enemies (e.g., predators, parasites, and pathogens), by genetic resistance in host 

plants, and by natural conditions or man-made environmental modifications (e.g., 

fallows, hedge rows, flooding) that interrupt reproductive cycles of pest species, 

including weeds. Natural pest control helps maintain the stability and diversity of 

ecosystems and reduces societal reliance on chemical pest control. 

• Pollination refers to the dependence of many plants on insects or other wild animals for 

sexual reproduction (i.e., transfer of pollen). Successful pollination contributes to the 

overall maintenance of both plant and animal diversity in an ecosystem. Pollination is 

related to species habitat because the availability of pollinators can be affected by the 

availability of their foraging and reproductive habitats. 

Process/Infrastructure 

• Microclimate control includes processes such as shading and wind breaking that 

provide local and regional temperature control. Microclimate control is essential to 

maintaining the structure and quality of many wildlife habitats. 

• Geomorphological control describes the services that maintain the physical integrity 

and structure of ecosystems and wildlife habitats. Specific processes include the 

following: organic production and export, sediment trapping, soil generation, flood 

control and desynchronization, storm surge protection, wave and wind buffering, 

shoreline anchoring, erosion control, and disturbance recovery. 

• Water supply includes the quantity and quality of groundwater and surface water, which 

influence both the amount and quality of available aquatic habitat and characteristics of 

terrestrial vegetation. For example, groundwater recharge protects aquifers from 

saltwater intrusion in coastal areas, which otherwise could alter the species composition 

of the local plant, and therefore animal, communities. Terrestrial animals also rely on 

water for basic life support functions. 

Page 50 
• Energy and nutrient exchange refers to processes that control the flow of energy, 

minerals, and nutrients (such as nitrogen, phosphorus, carbon, and sulfur). For example, 

photosynthesis by primary producers captures energy from the sun and converts 

inorganic carbon to organic carbon. Decomposition of dead biotic materials is essential 

to the provision of essential raw materials. These energy and nutrient exchange 

processes make energy and essential raw materials available to other organisms. 

• Purification of resources includes the retention and detoxification of pollutants as well 

as the removal of excess nutrients. The retention and detoxification of pollutants can 

reduce adverse effects on survival, growth, and reproduction in wildlife. The removal of 

excess nutrients by microorganisms can maintain the integrity of aquatic ecosystems by 

preventing algal blooms and anoxic conditions where they do not occur naturally. 

These neglected benefits can prove particularly useful in developing conceptual models for 

economic assessments of ecological benefits, as described in Chapter 4. 

3.4 TYPES OF ECOLOGICAL ASSESSMENTS 

Ecological assessment is a process used to evaluate changes to ecological resources resulting 

from natural or manmade events. Ecological assessments rely on the principles of ecology, 

discussed above, to identify, describe, and estimate the consequences of a change to any 

component(s) of an ecosystem. The changes may be biological (e.g., introduction of a nonnative 

predatory species), chemical (e.g., presence of a toxic chemical), or physical (e.g., loss of 

habitat). The ecosystem effects of the changes depend on those components of the ecosystem 

that are directly impacted and interactions of those components with the rest of the ecosystem. 

3.4.1 Assessment Models 

Prospective (stressor-driven) and Retrospective (effects-driven). Ecological risk/benefit 

assessments can be used to estimate the likelihood of future adverse effects or improvements 

(prospective) or to evaluate the likelihood that existing effects are caused by past exposure to 

stresses or removal of stresses (retrospective). Prospective risk assessments are stressor driven, 

and prospective benefit assessments are driven by proposed management options. Retrospective 

risk assessments are impact driven. For example, most watershed-level ecological risk 

assessments have been driven by observations of loss of water quality and degradation of aquatic 

communities, with the associated loss of recreational and fisheries values of the waters. 

Retrospective benefit assessments could be conducted to determine the efficacy of actions taken 

to help improve ecosystem condition. 

Individual-level ecological risk or benefit assessments are used only for endangered species. 

The ecological endpoints for assessing risks or benefits to such species include individual 

survivorship, growth, health, and reproductive success. This level of assessment ignores the rest 

of the species in an ecosystem and higher-level ecological entities. This approach can be useful 

in assessing threats or benefits to endangered species in the context of actions that would affect 

that species. Because individuals of endangered species are considered valuable, this approach 

generally looks for a threshold for individual-level effects rather than dose-response information. 

Page 51 
This level of assessment generally misses ecosystem resources and services that should be 

addressed in economic assessments of ecological benefits. 

Population-level assessments can be approached from the bottom up using data on individuallevel endpoints in or from the top down by modeling interactions among species in a community. 

The bottom-up approach uses individual-level endpoints, including mortality, growth, and 

reproductive success, to model population growth or sustainability. The top-down approach uses 

models of competition or predatory-prey relationships. Population-level assessment endpoints 

include the number of organisms that can be harvested on a sustainable basis, population density, 

and the probability of extinction. Where sustainable populations of a species are valued because 

of their direct use (e.g., consumption) by humans, a population-level assessment is key to an 

economic assessment of benefits. Information on life history strategies is important to 

population-level assessments. As Suter (1993) points out, populations of long-lived vertebrates 

(e.g., whales, seabirds) are more sensitive to changes in adult mortality than are shorter-lived 

species (e.g., quail, grasshoppers) that produce large numbers of offspring at each reproductive 

effort. The shorter-lived species can be more sensitive to short-term catastrophic events that 

coincide with and affect critical life stages (Suter, 1993). 

Many approaches have been developed to conduct population-level assessments, including 

modeling reproductive potential (e.g., using the “Leslie matrix”), aggregated models, and 

individual-based models (noted above) using logistic growth or age-structured population 

models (Suter, 1993, Gotelli, 1998). Aggregated models use aggregate components, such as 

population size or adults and juveniles, to assess population-level effects. These models are 

simplified versions of models of age-structured populations, where age is divided into one or two 

classes. 

A difficulty with modeling population growth or harvest potential is that the factors limiting a 

population (e.g., food resources or predation) often are difficult to identify or measure in the 

field (Suter, 1993). Most populations exhibit density-dependent mortality and growth, which 

complicates the modeling process. However, simple density-independent models have been used 

successfully by fish and wildlife managers where time horizons are short and expected changes 

are small (Suter, 1993). Gulland (1977) provides helpful discussions of surplus production 

models and various stock-recruitment models developed by fisheries biologists. 

Community-level assessments focus on the interactions among species in the community, 

including predator-prey and competitive relationships. These interactions can be much more 

sensitive to a stress than the individual-level endpoints (e.g., mortality, growth, and 

reproduction) used to estimate population-level effects. For example, chemical contaminants 

might impair the ability of a prey species to detect (e.g., sensory impairment) or escape from a 

predator (e.g., motor impairment) at levels that do not otherwise affect the growth and 

reproduction of the species. Many species interactions are of direct economic importance, such 

as the predatory behavior of biocontrol agents and the symbiosis between legumes and nitrogenfixing bacteria (Suter, 1993). Assessment endpoints at the community-level include community 

trophic structure and indices of community species composition (e.g., Index of Biotic Integrity, 

IBI). Difficulties with community-level assessments include selecting the interactions on which 

to focus and data availability. To the extent that an action or a stressor impacts the relationship 

of some species with their abiotic environment, ecosystem-level assessments can be needed. 

Page 52 
Ecosystem-level assessments use ecosystem properties, such as eutrophication, changes in 

biodiversity, and net productivity, as the assessment endpoints. They also can be used to predict 

changes in community or population-level endpoints (Suter, 1993). For example, an ecosystem 

model can be used to predict the effects of changes in populations at lower trophic levels on 

species at higher trophic levels and changes in community structure that result. Ecosystem-level 

endpoints are needed to assess changes in many of the beneficial ecosystem services (e.g., flood 

protection, water filtration, microclimate control). In general, ecosystem-level observations and 

data are needed to support stressor-response profiles for ecosystem-level responses. Relatively 

few laboratories run such tests on toxic chemicals. Stressor-response data for ecosystem services 

tends to be derived from field studies in which parameter measures (e.g., areal extent of wetland 

water recharge rate) are correlated with indicators of the ecosystem service (e.g., flood control). 

A key uncertainty in using ecosystem-level stressor-response data in prospective risk/benefit 

assessments is in extrapolating from the observed ecosystem to other ecosystems (Suter, 1993). 

The most common practice in ecological risk assessments at this time is to predict ecosystemlevel responses from the bottom up, extrapolating from lower-level responses to the ecosystem 

level. Because of functional and structural redundancies in ecosystems, significant effects can 

occur in single species without affecting ecosystem structure or function. An ecosystem cannot 

be more sensitive than its most sensitive component. Thus, it is a common practice in risk 

assessments to identify thresholds for adverse ecosystem effects based on individual-level effects 

on one of the most sensitive species. For example, EPA’s National Water Quality Criteria 

identify a threshold for adverse ecosystem-level effects as the best estimate (fiftieth percentile) 

of a concentration that would protect 95 percent of the species in the system (Stephan et al., 

1985). Use of individual-level effects to predict a threshold for ecosystem-level effects is not 

helpful, however, for the economic assessment of ecological benefits because it does not allow 

estimates of the magnitude of response, i.e., it does not use exposure-response data to predict the 

degree of change in response to an action of a specified magnitude. 

Landscape-level assessments are needed where watersheds are at issue or where the geographic 

distribution, connectivity, and diversity of different ecosystems or habitats over a large 

geographic area can be affected by a proposed action. MacArthur and Wilson’s (1963, 1967) 

theory of island biogeography states that the low species diversity characteristic of oceanic 

islands reflects a dynamic equilibrium between rates of extinction and rates of colonization of 

individual species’ populations. This theory has been used to assess the effects of habitat loss 

and fragmentation in terrestrial environments on animal species, including neotropical migrant 

songbirds, wolves, turtles, and many others. Landscape-level assessments are also needed to 

assess “edge effects,” such as higher bird nest predation rates by blue jays and crows at the edge 

of forests rather than in the interior of forests (Terborgh, 1989; Wilcove, 1985). Nest parasitism 

by brown-headed cowbirds also occurs at the edge of forested habitats (Terborgh, 1989). Edge 

effects can penetrate several hundred meters into forested habitats. Thus landscape-level 

measures such as the ratio of the length of forest edge to the area of the forest interior can be 

useful in predicting population-level responses of valued species. 

Page 53 
3.4.2 Standardized Approaches to Ecological Assessments 

Ecological assessments have been conducted under a variety of statutes by federal and other 

agencies for some decades. Standardized approaches and nomenclature for those assessments 

have developed somewhat independently between the different agencies and offices. 

Differences among the approaches can be attributed to different statutory requirements, the type 

of information generally available at the start of an assessment, and the experience of the agency 

or office responsible for the assessment. The remainder of this section briefly describes the 

primary different types of ecological assessments and explains why EPA’s (1998) Guidelines for 

Ecological Risk Assessment was used as the starting point for development of this Framework. 

Types of Standardized Approaches 

A variety of different “standardized” approaches to ecological risk and impact assessments have 

developed over the past few decades in response to various legal mandates administered by 

different agencies. Key attributes of some of the more well-recognized approaches are listed in 

Exhibit 11. 

Ecological Risk Assessment. As defined by EPA (1998), an ecological risk assessment is a 

process that evaluates the likelihood that adverse ecological effects may occur or are occurring 

as a result of exposure to one or more stressors. Ecological effects can be evaluated both 

qualitatively or quantitatively in terms of structural and functional changes at one or more levels 

of biological organization. In EPA’s paradigm, the part of the ecosystem that is affected by the 

change is called a “receptor(s)” and is usually a structural component. The natural or 

anthropogenic (i.e., manmade) event causing the effects is called a “stressor.” Stressors can be 

chemical, physical, or biological. The “effects” of the stressor include direct changes to the 

receptor(s) as well as indirect changes to other structural or functional components that are 

affected through the interconnections that define the ecosystem (e.g., energy flows through the 

food web). 

Environmental Assessment (EA). An EA is frequently required under the National 

Environmental Policy Act (NEPA), prior to, and in some cases, in lieu of, the preparation of an 

Environment Impact Statement. EAs are concise documents prepared on a case-by-case basis by 

government agencies. They describe the environmental impacts of a proposed government 

action, provide a listing of agencies or persons consulted, and discuss possible alternative 

actions. There must also be an evaluation of the probable cumulative, long-term environmental 

effects including any beneficial impacts. 

Environmental Impact Statement (EIS). An EIS is a type of assessment that attempts to 

reveal the consequences of a proposed action as an aid to governmental decisionmaking. In the 

United States, federal agencies are required by NEPA to prepare an EIS for any “major federal 

action.” Similar requirements exist for some states as well as for a few other nations. The scope 

Page 54 
Exhibit 11 

Standardized Approaches to Ecological Risk Assessment 

Standard Name for 

Assessment 

Statute Key Attributes Process 

Ecological Risk Assessment various, e.g. prospective/stressor driven, traditionally tiered 

(ERA): national-level RCRA, CAA chemical, now being adapted to other 

stressors 

ERA: regional, landscape, or various, including prospective/stressor or impact driven, tiered 

watershed level CWA usually chemical 

ERA: site-specific various, e.g., prospective/stressor driven, usually tiered 

CERCLA, RCRA chemical 

Environmental Assessment 

(EA) 

NEPA prospectiv/stressor driven, all types screening 

Environmental Impact NEPA prospective/stressor driven, predominantly refined 

Assessment (EIA) physical stressors 

Habitat Assessment (and NEPA and land prospective/stressor driven, land refined 

Habitat Suitability Index) management management changes 

Hazard Assessment various, e.g., prospective/stressor driven, traditionally screening, 

TSCA and FIFRA chemicals and tiered 

Natural Resource Damage CERCLA retrospective/impact driven, chemicals refined 

Assessment (NRDA) only 

Acronyms for Statutes: CAA - Clean Air Act; CWA - Clean Water Act; CERCLA - Comprehensive Environmental 

Response, Compensation, and Liability Act; FIFRA - Federal Insecticide, Fungicide, and Rodenticide Act; RCRA ­

Resource Conservation and Recovery Act; NEPA - National Environmental Policy Act; TSCA - Toxic Substances 

Control Act. 

and content of an EIS depend on the type of activity under consideration. An EIS is required to 

predict any or all future effects on the environment. Consequently, it devotes considerably more 

attention to identifying the full range of affected environmental components, defining the 

geographic and temporal changes, and identifying secondary and tertiary effects than an EA. 

NEPA explicitly states a policy of preserving the quality-of-life benefits of natural areas and 

resources for future generations and evaluating cumulative impacts of activities over time. 

Habitat Assessment. Habitat assessments evaluate the suitability of a local habitat to support a 

given species. The most well known example is the U.S. Fish and Wildlife Service’s Habitat 

Evaluation Procedure (HEP). HEP provides a framework for determining habitat quality for 

specific fish and wildlife species by quantifying many characteristics of the environment, 

including physical, chemical, and biological characteristics (Scodari, 1992). The relationships of 

different values of those characteristics to the species’ population densities and reproductive 

success (implied suitability) have been developed from previous field studies. Use attainability 

analyses performed under the Clean Water Act are also considered habitat assessments. They 

determine what uses of a water body are attainable (e.g., swimming, fishing, water supply), the 

Page 55 
extent to which pollution is impacting these uses, and the necessary pollution control measures 

that are needed. Use attainability analyses must consider habitat limitations such as frequency of 

low tides, natural water quality, and physical structure of the habitat. 

Hazard Assessment. Hazard assessments determine the existence of a hazard. This type of 

assessment identifies the types of effects a particular stressor might have on different groups of 

organisms based on experimental exposures of organisms to the stresses. The hazard assessment 

helps to identify particularly sensitive groups of organisms (or functions), which in turn affect 

the selection of assessment endpoints. The phrase hazard assessment has been used to describe a 

comparison of the magnitude of expected levels of stress in the environment to thresholds of 

effect in groups of organisms (or functions). Currently, that activity is more appropriately called 

a screening-level ecological risk assessment. 

Natural Resource Damage Assessment (NRDA). Natural Resource Damage Assessments are 

retrospective assessments that address both ecological and economic damages. Standard 

methodologies promulgated by the Department of the Interior (DOI) and the National Oceanic 

and Atmospheric Administration (NOAA) require an assessment of injury to an ecological 

resource and an evaluation of the economic damages. In an NRDA, federal or state officials, 

acting as trustees for natural resources, can seek compensation from responsible parties under the 

Oil Pollution Act, CERCLA, and other statutes for damages to natural resources (e.g., loss of 

shellfish beds) caused by releases of oil and other toxic materials. Trustees have used NRDA 

regulations to seek monetary compensation for natural resource injuries associated with 

accidental releases, such as the Exxon Valdez oil spill. A NRDA may be conducted at a 

Superfund site at the discretion of natural resource trustees. An injury assessment, which 

documents the adverse effects associated with a release, is the basis for the NRDA. An injury 

assessment is basically a retrospective risk assessment to link injuries to particular contaminant 

sources. 

Selection of Approach for this Framework 

Of the various standardized approaches noted above, EPA’s (1998) Guidelines for Ecological 

Risk Assessment is the most general and flexible, because the Guidelines document was designed 

to encompass the broad range of statutory requirements that different EPA Offices administer. 

Although historically, EPA has focused on assessments of chemical contaminants in the 

environment, the Agency intentionally included information on risk assessment for physical and 

biological stressors as well to further broaden the scope and utility of the Guidelines within and 

beyond the Agency. That Guidelines document, for example, is general enough to be adapted to 

benefit analyses (U.S. EPA, 1998). That Guidelines document was developed by EPA’s Risk 

Assessment Forum over a period of years with substantial input from EPA and other federal 

agencies. Drafts of the proposed Guidelines received extensive scientific peer review and 

interagency committee review. For these reasons, the framework described by EPA’s (1998) 

Guidelines for Ecological Risk Assessment was used as the starting point for building the 

framework for the economic assessment of ecological benefits proposed in this document. The 

adaptation of EPA’s (1998) Guidelines for Ecological Risk Assessment to the ecological benefits 

assessment addressed in this document is discussed in the next chapter. 

Page 56 
The next chapter goes into more detail on conducting an ecological risk/benefit analysis. The 

purpose of Chapter 4 is to help the economist better understand the scientific framework for 

analysis and type of information that may be generated through an ecological assessment. 

Improved understanding of the ecological risk assessment process will facilitate communication 

between economists and ecologists during planning of ecological assessments and thereby 

increase the utility of assessment results for economic analyses. 

References and Further Reading 

Allen, T.F. and Hoekstra, T.W. 1992. Toward a Unified Ecology. Complexity in Ecological 

Systems Series. New York, NY: Columbia University Press. 

Briand, F., and Cohen, J.E. 1987. Environmental correlates of food chain length. Science 238: 

956-960. 

Brown, J.H. and Lomolino, M.V. 1998. Biogeography. 2nd Ed. Sunderland, MA: Sinauer 

Associates, Inc. Publishers. 

Costanza, R., R. d’Arge, R. de Groot, S. Farber, M. Grasso, B. Hannon, K. Limburg, S. Naeem, 

R.V. O’Neill, J. Paruelo, R.G. Raskin, P. Sutton, and M. van den Belt. 1988. The value of the 

world’s ecosystem services and natural capital. Ecological Economics 25(1):3-15. 

Daily, G.C. (Ed). 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. 

Washington, DC: Island Press. 

Daily, G.C., S. Alexander, P.R. Ehrlich, L. Goulder, J. Lubchenco, P.A. Matson, H.A. Mooney, 

S. Postel, S. H. Shneider, D. Tilman, and G.M. Woodwell. 1997. Ecosystem services: benefits 

supplied to human societies by natural ecosystems. Ecological Society of America, Issues in 

Ecology, Number 2, Spring 1997. 

Gotelli, N.J. 1998. A Primer of Ecology, 3rd Ed. Sunderland, MA: Sinauer Associates Inc. 

Gulland, J.A. 1977. Fish Population Dynamics. London, UK: John Wiley & Sons. 

Hairston, N.G. Jr., Hairston, N.G. Sr. 1993. Cause-effect relationships in energy flow, trophic 

structure, and interspecific interactions. Am. Nat. 142: 379-411. 

Hall, S. J.; Raffaelli, D. G. 1993. Food web: theory and reality. In: Begon, M.; Fritter, A. H., 

eds. Advances in Ecological Research, Vol. 24. San Diego, CA: Academic Press; pp. 187-239. 

Likens, G. 1992. An Ecosystem Approach: Its Use and Abuse. Excellence in Ecology, Book 3, 

Ecology Institute, Oldendorf/Luhe, Germany.) 

MacArthur, R.H., and Wilson, E.O. 1963. An equilibrium theory of insular biogeography. 

Evolution 17: 373-387. 

Page 57 
MacArthur, R.H., and Wilson, E.O. 1967. The Theory of Island Biogeography. Princeton, NJ: 

Princeton University Press. 

Margalef, R. 1968. Perspectives in Ecological Theory. Chicago, IL: University of Chicago 

Press. 

Norse, E. 1990. Threats to Biological Diversity in the United States. Report prepared for the 

U.S. EPA, Washington, DC, by Industrial Economics, Contract No. 68-W8-0038, Work 

Assignment 115.

Odum, E.P., in collaboration with H.T.Odum. 1959. Fundamentals of Ecology. Philadelphia,

PA: Saunders.

Pimentel, D. 1988. Economic benefits of natural biota. Ecological Economics 25(1):45-47.

Pimm, S.L. 1980. Properties of food webs. Ecology 61: 219-225.

Pimm, S.L. 1982. Food Webs. New York, NY: Chapman and Hall.

Principe, P.P. 1995. Ecological benefits assessment: A policy-oriented alternative to regional

ecological risk assessment. Human and Ecological Risk Assessment 1(4):423-435.

Scodari, P. 1992. Wetland Protection Benefits. Draft Report. Prepared for U.S. EPA, Office of

Policy, Planning, and Evaluation under Grant No. CR-817553-01. October.

Stephan, C.E., Mount, D.I., Hanson, D.J., Gentile, J.H., Chapman, G.A., and Brungs. W.A. 

1985. Guidelines for Deriving Numeric National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses. Duluth, Minnesota: U.S. EPA. NTIS No. PB85-227049. 

Suter, G.W. II. 1989. “Ecological Endpoints.” in Warren-Hicks, W., B.R. Parkhurst, and S.S. 

Baker, Jr., eds. Ecological Assessment of Hazardous Waste Sites: A Field and Laboratory 

Reference Document. EPA/600/3-89/013. Corvallis Environmental Research Laboratory, 

Oregon. 

Suter, G.W. II. 1993. Ecological Risk Assessment. Boca Raton, FL: Lewis Publishers. 

Suter, G. W., Efroymson, R.A., Sample, B.E., Jones, D.S. 2000. Ecological Risk Assessment for

Contaminated Sites. Boca Raton, FL: Lewis Publishers.

Terborgh, J. 1989. Where Have All the Birds Gone? Princeton, NJ: Princeton University Press.

U.S. EPA. 1992a. Framework for Ecological Risk Assessment Washington, DC: U.S. EPA, 

Risk Assessment Forum. EPA/630/R-92/001. February. 

U.S. EPA. 1992b. Biological Populations as Indicators of Environmental Change, 

EPA/230/R-92/011. Washington, DC: Office of Policy Planning and Evaluation. 

Page 58 
U.S. EPA. 1993. Habitat Evaluation: Guidance for the Review of Environmental Impact 

Assessment Documents. Prepared by Dynamac Corporation for the Office of Federal Activities 

under U.S. EPA Contract No. 68-C0-0070. January. 

U.S. EPA. 1994a. Managing Ecological Risks at EPA: Issues and Recommendations for 

Progress. Prepared by M.E. Troyer and M.S. Brody. Washington, DC: U.S. EPA. EPA/600/R94/183. 

U.S. EPA. 1994b. Toward a Place-Driven Approach: The Edgewater Consensus on an EPA 

Strategy for Ecosystem Protection. Ecosystem Protection Workgroup. Washington, DC: U.S. 

EPA. March 15 Draft. 

U.S. EPA. 1994c. Background for NEPA Reviewers: Grazing on Federal Lands. Prepared by 

Science Applications International Corporation under U.S. EPA Contract No. 68-C8-0066. 

February. 

U.S. EPA. 1997. Priorities for Ecological Protection: An Initial List and Discussion Document 

for EPA. Washington, DC: U.S. EPA. EPA/600/S-97/002. 

U.S. EPA. 1998. Guidelines for Ecological Risk Assessment. Washington, DC: U.S. EPA. 

EPA/630/R-95/002B. 

U.S. EPA. 2000. Assessing the Neglected Ecological Benefits of Watershed Management 

Practices: A Resource Book. Prepared for the Assessment and Watershed Protection Division, 

Office of Water by ICF Consulting. Washington, DC: U.S. EPA. April. 

Wilcove, D.S. 1985. Nest predation in forest tracts and the decline of migratory songbirds. 

Ecology 66: 1211-1214. 

Wilson, E.O. and Bossert, W.H. 1971. A Primer of Population Biology. Stamford, CT: Sinauer 

Associates Inc. Publishers. 

Page 59 
4.0 ECOLOGICAL RISK/BENEFIT ASSESSMENT 

As indicated in Chapter 3, the framework used for an ecological benefits assessment in this 

document is based on EPA’s (1998) Guidelines for Ecological Risk Assessment. This section 

describes the adaptation of that framework to build the proposed framework for the economic 

assessment of ecological benefits. 

4.1 OVERVIEW OF EPA’S GUIDELINES FOR ECOLOGICAL RISK ASSESSMENT 

An ecological risk assessment determines the likelihood, potential nature, and magnitude of an 

adverse ecological effect resulting from exposure to a stressor (U.S. EPA, 1998). Some 

examples of ecological stressors are listed below: 

Physical Biological 

C Erosion C Disease-causing organisms (Pfiesteria, C Heat diatoms) C Turbidity C Genetically-engineered microorganisms C Impoundments C Non-native species (kudzu, zebra C Habitat alterations mussels)

Chemical 

C Hazardous substances (e.g., pesticides, industrial wastes) 

C Salinity 

C Air pollutants (CO, NOx, ozone, hazardous air pollutants ) 

The description of potential ecological effects should include magnitude, duration, spatial

distribution, time to recovery, and other relevant parameters. As indicated in Chapter 3,

ecological risk assessments may be predictive (i.e., estimate the probability and magnitude of

future ecological changes in response to a given stressor), or they may be retrospective, (i.e., 

assess the probability that a past event caused this present problem). A predictive benefits

assessment may take the form of modeling the effects of an activity (e.g., removal of a stressor),

such as the effects of reducing atmospheric nitrogen levels on the Chesapeake Bay. Such

assessments depend on applying previously collected data from similar events and ecosystems to

a new situation. In a retrospective ecological benefits assessment, an effect may be well defined,

such as reduced sedimentation, but the potential of other environmental changes contributing to

the improvement must be considered. 

Benefits can be expressed quantitatively (e.g., a probability, such as an 80 percent chance that a

population will not go extinct in the next 100 years) and/or qualitatively (e.g., low, medium, or

high). In addition, the uncertainty associated with the probability needs to be addressed, either

quantitatively or qualitatively.

Page 60 
Steps in an Ecological Benefits Assessment 

As illustrated in Exhibit 2 of Chapter 2, an ecological benefits assessment starts with planning. 

During planning, the environmental decision makers, risk assessors, economists, and others 

determine the need for and scope of the ecological benefits assessment. It is during this stage 

that societal and political issues are considered, and a key to success is determining the 

involvement of different individuals in planning the risk assessment. Participants in the planning 

phase may include risk assessors (including scientists), risk managers (e.g., government 

regulators), economists, and, if appropriate, interested outside parties (e.g., environmental and 

industry groups, those whose land may be affected by risk assessment decisions, state and/or 

local government officials, etc.). Collaborative planning can help foster a consensus on which 

ecological benefits are most valuable to the stakeholders and the goals, scope, and timing of the 

ecological benefits assessment. 

Stakeholder Involvement in Planning 

Waquoit Bay provides an excellent example of how stakeholder involvement can be 

instrumental in developing management goals for an ecological risk or benefits 

assessment. It has been generally agreed by all involved parties that the Bay is changing 

— eelgrass is disappearing and is being replaced by thick mats of macro algae, fish kills 

are occurring, and scallops have disappeared. Something must be done to prevent further 

degradation and restore what has already been damaged. Three steps were used to 

develop management goals for Waquoit Bay: 

• A public meeting of all stakeholders; 

• An evaluation of written goals by organizations having jurisdiction or an interest 

in the ecology of the watershed; and 

• A meeting of members of these organizations to review and approve the 

management goals. 

The public meeting was advertised in local newspapers. The meeting was designed to 

determine what the public viewed as valuable in the bay and what the main stressors 

were on these values. The participants found the bay to be valuable for a number of 

reasons including open space, scenic views, flyways for waterfowl, shellfishing, 

navigation, wildlife, and human serenity. Stressors were many. They included physical, 

chemical, and biological impacts to the bay such as the introduction of non-native 

species, man-made noise, fertilizers, ignorant tourists, habitat loss, and boat wake 

disturbance. 

Numerous governmental (federal, state, and local) and non-governmental organizations 

were involved in the review and approval of the management goals. The groups 

involved in developing these goals are considered the risk management team for the 

watershed and will be principally responsible for implementing the management plan in 

Waquoit Bay. 

Page 61 
Planning might be considered complete once the following objectives have been met. 

• Objectives of the risk assessment have been defined (including criteria for success); 

• Goals for ecological values have been established; 

• The range of options under consideration has been developed; 

• Focus and scope of the assessment have been agreed upon; and 

• Resources to conduct the assessment have been provided (U.S. EPA, 1998). 

As described in Chapter 2, EPA defines three phases for conducting ecological risk assessments 

(U.S. EPA, 1998) that can be adapted for ecological benefits assessments: 

• Phase 1 — Problem Formulation; 

• Phase 2 — Analysis (exposure assessment and ecological effects characterization); and 

• Phase 3 — Risk Characterization. 

A somewhat simplified diagram of EPA’s risk assessment process is provided in Exhibit 12. 

Readers interested in EPA’s (1998) Guidelines as they pertain to risk instead of benefit 

assessment are referred to that document. The remainder of this Chapter describes how those 

Guidelines have been adapted to the proposed framework for economic assessments of 

ecological benefits. 

4.2 PHASE I: PROBLEM FORMULATION 

Problem formulation in 

an ecological benefits 

assessment is the process 

for generating and 

evaluating preliminary 

hypotheses about what 

benefits might accrue 

from a proposed action 

and its alternatives. It 

provides the foundation 

for the entire economic 

assessment of ecological 

benefits. Problem 

l ic 

l i

Identify Endpoints and Linkages Between Them 

Ecologica

Prioritize Endpoints and Select Valuation Techniques 

Develop Analysis Plans, Ensuring Analytical Compatibility 

Develop Conceptual Models Econom

Ecological Analysis Plan Economic Analysis Plan 

Integrate Available Information 

Prob em Formulat on Phase 

Page 62 
formulation involves the development of three products: assessment endpoints, a conceptual 

model, and an analysis plan. 

Page 63
Exhibit 12 

Ecological Risk Assessment Framework 

of of 

/ 

) 

Characterization 

Exposure 

Characterization 

Ecological 

Effects 

Conceptual 

Model 

Assessment 

Endpoints Analysis 

Plan 

PROBLEM FORMATION 

Planning 

(Risk Assessor

Risk Manager 

Dialogue

Risk 

Estimation 

Risk 

Description 

RISK 

CHARACTERIZATION 

ANALYSIS 

Communicating Results 

to the Risk Manager 

Risk Management 

Page 64
The development of these products requires several activities: 

• Integrating available information; 

• Using an iterative process and coordination between the ecological assessment team and 

the economic analysts to develop a joint conceptual model that links ecological 

assessment endpoint to economic benefit endpoints; 

• Prioritizing endpoints jointly with the economists; and 

• Developing the analysis plan, ensuring analytic compatibility with the economic 

assessment analysis plan. 

4.2.1 Conceptual Model 

As described in Chapter 2, the ecologists develop their preliminary conceptual model based on 

the cascade of ecological effects expected from the action itself to direct effects and indirect 

effects. The conceptual models presented in Chapter 2 are examples of preliminary models in 

which little attention has been given as yet to what might be considered as assessment endpoints. 

The ecologists and the economists work together in an iterative process refining the model to 

ensure comprehensive coverage and appropriate linkages between ecological endpoints and 

economic benefit endpoints. 

In contrast to the goals of many site-specific ecological risk assessments (e.g., to protect an 

aquatic community by protecting the most sensitive species in the community), a goal of the 

economic assessment of ecological benefits is to capture as many of the potential benefits that 

might result from an action as possible. The emphasis in the benefits conceptual model is to 

identify the full cascade of ecological effects that might result. In developing the conceptual 

model of likely ecological benefits, the framework described in EPA’s (2000) report Assessing 

the Neglected Ecological Benefits of Watershed Management Practices can be helpful. The 

report discusses linkages between watershed management practices and the neglected benefits 

(see Chapter 3), including a general estimate of the strength of the linkages. Exhibit 13 below 

provides an example of the neglected ecological benefits that can accrue from the forestry 

management practices of revegetation and forest regeneration. Exhibit 13 also provides specific 

examples of benefit endpoints within each category of neglected benefits. 

The conceptual model is accompanied by hypotheses about how the initial action causes both 

direct and indirect effects. The complexity of the conceptual model depends on the complexity 

of the problem (e.g., number and types of stressors, number of assessment endpoints, nature of 

effects, and characteristics of the ecosystem). 

A conceptual model diagram (see Exhibit 14) is a useful way to visually express the 

relationships described by the benefit hypotheses. Conceptual model diagrams can communicate 

important exposure pathways in a clear and concise way, facilitating the coordination between 

ecologists and economists in problem formulation. These diagrams and hypotheses also are 

useful tools to aid communication with the environmental decisionmakers and interested parties. 

Page 65 
The number of relationships that can be depicted in one flow diagram depends on the 

comprehensiveness of each relationship. The more comprehensive the relationship, the fewer 

relationships that can be shown with clarity in one diagram, thus separate diagrams may be 

required. There is no set configuration for conceptual model diagrams. 

In developing the conceptual model, the ecologists can consider each proposed action in relation 

to each category of neglected benefits, trying to identify and specify ecological benefits in that 

category that might result from the proposed action. Using the list of neglected benefits should 

help to ensure that the ecological conceptual model is comprehensive and that as many 

ecological benefits of a proposed action as possible are identified for the economic assessment. 

Exhibit 14, illustrates where use of “neglected” ecological benefits in the conceptual model 

helps to link the action to important and quantifiable ecological benefits. 

Exhibit 13 

Examples of Benefits Associated with Forestry Practice of Revegetation 

and Forest Regeneration (from U.S. EPA, 2000) 

Neglected Benefit Example(s) of Benefit(s) Provided by Watershed 

Category Management Practice(s) 

Species Habitat • increase in species habitat provided by new trees 

Biotic Productivity • contribution of new trees to primary productivity 

• maintenance of aquatic plant productivity through decreased water column turbidity 

and increased light penetration 

Biodiversity • maintenance of natural species assemblages by planting native species 

• maintenance of or increase in biodiversity by increasing available forest habitat for 

threatened or endangered species 

Food Chain Support • maintenance of species and population balance within food web 

Microclimate Control • increase in stream shading due to new trees in riparian areas 

• maintenance of wind speed 

Geomorphological • prevention of landslides and erosion by stabilizing soil 

Control 

Water Supply • decrease in soil erosion that could otherwise increase water turbidity and nutrient 

loading 

Energy and Nutrient • preservation of soil microbes by maintaining leaf litter levels 

Exchange 

Purification of • preservation maintenance of decomposition organisms in leaf litter and soil 

Resources • interception and retention of airborne pollutants 

Page 66 
Exhibit 14 

Conceptual Model for an Economic Assessment of Ecological Benefits 

Improved Local Septic Systems 

Reduced Nutrient Loading 

Improved Aquatic Habitat Improved Wetland Structure and 

Function 

Stressor Reduction > 

“Neglected” Ecological 

Ecological Assessment 

“Neglected” Ecological 

Ecological Assessment 

Reduced 

Property Losses 

Improved Bird 

Watching 

Increased Fish 

Landings 

Increased 

Recreational 

Swimming 

Increased Use 

by Migratory 

Birds * 

Increased 

Shorebird 

Populations * Improved Water 

Quality 

Improved Storm 

Protection 

Improved Water 

Filtration 

Increased Emergent 

Vegetation Density and 

Areal Extent 

Increased Fish and 

Shellfish Populations 

Reduced Eutrophication in Local Waters 

Action > 

Primary Effect > 

Benefits > 

Secondary Effects; 

Endpoints > 

Benefits > 

Tertiary Effects; 

Endpoints > 

Economic 

Endpoints > 

Key to Symbols: Action 

Stressor 

Ecological 

Assessment 

Endpoints 

Other 

Ecological 

Effects 

Economic 

Endpoints 

* Indicates assessment endpoints that might not be needed for an ecological risk assessment, but are added to assist 

the economic benefit analysis. 

Page 67 
4.2.2 Assessment Endpoints 

The conceptual model of the ecological benefits assessment should include well-defined 

assessment endpoints that will be linked to the economic benefit endpoints. In the conceptual 

models in Chapter 2, a series of vague endpoints were linked, with “improved wetland structure 

and function” leading to improved water filtration and storm protection. However, the concept 

of improved wetland structure and function is too vague to allow quantitative links to those 

benefits. Thus, an assessment endpoint specified as “increased emergent vegetation density and 

areal extent” was inserted to allow quantification. 

Assessment endpoints may be defined for both structural and functional aspects of an ecological 

resource at any level of organization ranging from a single individual to an entire landscape. 

Exhibit 15 provides examples of assessment endpoints at different levels of biological 

organization. Most benefit endpoints, however, are likely to be defined at the population level or 

higher. This is true because a population is the lowest level of biological organization that can 

be meaningfully protected (Suter, 1993). An effect on one or several individuals will not 

necessarily result in significant population changes. Exceptions to this premise are threatened 

and endangered species for which each individual is valuable to the survival of the population. 

Exhibit 15 

Examples of Assessment Endpoints 

Level of Organization Ecological Entity Assessment Endpoints 

Individual bald eagle survivorship, reproductive success 

Population fish populations population density, population growth rate 

Community aquatic 

communities 

survival, development, and reproduction of 

fish, aquatic invertebrates, and plants 

Ecosystem lake eutrophication, nutrient flux 

Landscape habitat connectivity, ratio of length of edge to area 

of interior, patch size 

To assess interactions, it is important to use the most appropriate biological level of organization 

at which the interactions can be observed. For example, community-level effects can be 

obscured at the ecosystem-level of observation, some ecological changes can only be effectively 

evaluated from a landscape perspective, and so on. Assessment using a population -, 

community-, ecosystem-, watershed-, or landscape-level assessment may be most appropriate 

depending on the situation. For bioaccumulative compounds, exposure of the populations must 

be traced through the food web in order to understand the full magnitude of potential impacts to 

the community and ecosystem. The selection is based on the conceptual model on which the 

assessment is based. The conceptual model depicts the endpoints selected for assessment and 

the interactions (or linkages) among endpoints. 

Page 68 
Both structural and functional changes should be considered in a benefits assessment. 

Improvements of a specific habitat, such as an increase in wetland acreage resulting from the 

removal of a barrier to water flow, might yield both benefits that are both structural and 

functional in nature. In Exhibit 14, the conceptual model indicates that increased emergent 

vegetation and areal extent can provide improved water filtration services and improved storm 

protection because of the structural attributes of the emergent vegetation. 

Criteria for Prioritizing Assessment Endpoints 

As indicated in Chapter 2, there are both ecological and economic criteria that will be important 

in prioritizing the potential assessment/benefit endpoints for quantitative assessment. This 

section focuses on the ecological criteria. 

By addressing the needs for ecological 

Example of High Priority Assessment assessments under a variety of different 

programs, the Agency has identified some Based on EPA’s 1998 Criteria

general criteria to assist in selecting and 

prioritizing assessment endpoings. EPA’s Salmon abundance would rank high among 

the possible assessment endpoints for (1998) Guidelines for Ecological Risk 

Assessment evaluating the benefits of not constructing a specify three criteria by which to 

hydroelectric dam on a river in the Pacific select assessment endpoints in a risk 

Northwest. Salmon have ecological assessment: (1) have ecological relevance, (2) 

be susceptible to the stressor, and (3) and be relevance, because they are a food source

for many aquatic and terrestrial species, and relevant to the management goals (see text 

box). These criteria are particularly useful in they eat many aquatic invertebrates. 

Salmon are also sensitive to changes in assessments of the risk of adverse ecological 

sedimentation, water temperature, and effects. For an ecological benefits 

substrate pebble size. Most importantly, assessment, additional considerations and 

salmon are valued by society as a source of criteria for identifying high priority 

food, for recreational fishing, and for their ecosystems and ecological components can 

be useful. As described in Chapter 3, EPA ceremonial and symbolic significance to 

Native Americans. (1997a) has proposed four criteria for 

prioritizing ecological entities to be protected 

in its discussion document for Priorities for 

Ecological Protection: mandated protection, other societal values, rare or under threat, and 

ecological significance. For a benefits assessment, it is more appropriate to consider these 

attributes as prioritization criteria, instead of selection criteria. The remainder of this section 

discusses EPA’s 1998 criteria as they might apply to an ecological benefits assessment. 

Ecological relevance. The ecological relevance of an endpoint refers to its importance in 

relation to other components of a specific community or ecosystem. For example, if the change 

in abundance of certain benthic invertebrate species can affect the abundance or productivity of 

fish in the community, the benthic (i.e., bottom-dwelling) invertebrates are ecologically 

important. As another example, honeybees are ecologically significant in prairies because they 

pollinate many plants. Abundance or age structure of a certain population of fish may be 

selected as an assessment endpoint because of the critical role the species plays in maintaining 

the functional integrity of the ecosystem (i.e., top consumer that exerts control over lower 

Page 69 
trophic levels in the food web). Effects on primary producers, such as green algae in a lake, may 

be critical to higher trophic levels, such as insect larvae, that feed upon the algae, and fish, that 

feed upon the insect larvae. 

Susceptibility to the ecological stressor. For an ecological risk assessment, EPA’s (1998) 

Guidelines indicate that an assessment endpoint must be based on a link between a susceptible 

ecological component (or receptor) and exposure to an ecological stressor. Susceptibility is 

based on the sensitivity of the ecological receptor to the stressor and attributes of the life history 

of the receptor that might influence the likelihood (and magnitude) of exposure. Some 

organisms are more sensitive to stress at certain times in their life cycle such as during molting 

or during seed germination. In conducting an ecological risk assessment, it is desirable to 

determine effects on sensitive species and effects during sensitive life stages. For an ecological 

benefits assessment that evaluates the benefits of removing stressors from the environment, this 

criterion will be useful in selecting assessment endpoints. 

In a benefits assessment, the susceptibility of the different types of receptors included in the 

conceptual diagram need to be considered. In addition, the potential for some changes to be 

adverse while others are beneficial needs to be considered. For example, application of a 

pesticide might kill the target organisms such as mosquitos in a stagnant pond (primary effect) 

but it might also cause a decrease in the dragonfly population that feeds upon the mosquitos as a 

primary food source (the dragonflies may starve or leave the area seeking food elsewhere). 

Another factor to consider is that multiple stressors can increase the sensitivity of ecological 

components to any given stressor. A seal with a virus may be weakened and thus become easier 

prey for a shark. 

Relevance to management goals. An ecological risk assessment is most useful when the 

assessment endpoints are related to an ecological component or process that is valued by both 

the public and decisionmakers such as clean air in parklands. An ecological benefits assessment 

also requires that the economic benefit endpoints can be clearly linked to ecological endpoints. 

Ecological benefit assessments should include those ecological components that are used directly 

by humans, such as sport fish, groundwater, or timber land. Ecological benefit endpoints might 

also reflect those components and processes that indirectly benefit humans, such as water 

filtration, climate control, or flood protection. What is actually measured might be different 

from what is important to the economic valuation study, but the relationship between the two 

elements should be clearly defined. It also is important to remember that in many cases, 

selection of an assessment endpoint is constrained by an environmental law, such as the Clean 

Air Act, or a policy goal, such as pollution prevention via water permitting. 

Although not a specific criterion for selecting assessment endpoints, it is important that changes 

to the assessment endpoint can be predicted or measured, particularly if a stressor-response 

Page 70 
Linking the Assessment Endpoints to the Management Goals in Waquoit Bay 

The goal (or expected benefits) of the Waquoit Bay watershed management plan is to 

reestablish and maintain water quality and habitat conditions in the Bay to support diverse 

native fish and shellfish populations as well as reverse degradation of ecological resources 

in the watershed. One way to help accomplish this is to reestablish viable eelgrass beds and 

associated aquatic communities in the Bay. Therefore, an assessment endpoint was eelgrass 

abundance and distribution. Eelgrass is a rooted plant in the shallows of the Bay that 

decreases erosion and increases sedimentation, which in turn, provides food and habitat for a 

variety of marine organisms, such as juvenile scallops, invertebrates, and forage fish. 

Eelgrass is a good assessment endpoint because it has great ecological significance (i.e., it 

provides habitat for fish and shellfish.). The disappearance of eelgrass might have resulted 

from reduced light due to shading by algal blooms and turbidity from suspended sediments. 

Stressed eelgrass beds are also more susceptible to disease from slime mold. In addition, its 

distribution and acreage covered can be readily measured. 

Scallop abundance in the Bay is not a good assessment endpoint. Although it might have 

societal value, it would be difficult to assess whether changes in abundance resulted from 

natural variability or the effects of the stressor. A qualitative discussion of this endpoint 

might be useful, however, for benefits analysis. 

relationship can be established. If the assessment endpoint cannot be measured directly, 

appropriate surrogate components or qualitative values will need to be identified as well as 

methods for extrapolating effects to the assessment endpoints. For example, in many cases, a 

stressor-response relationship could be impossible to quantify even though the existence of a 

stressor-effect relationship is well established. This is the case when an affected population 

cannot be tested, such as an endangered species, or a situation where a synergist is involved. In 

these situations, it may be appropriate to simply indicate that an effect has been observed without 

indicating the intensity of the stressor. 

Combining Ecological and Economic Criteria for Prioritizing Endpoints 

In an economic assessment of ecological benefits, ecological selection or prioritization criteria 

alone are insufficient for selecting the ecological assessment endpoints that will be examined 

quantitatively. In addition, economic criteria, as described in Chapter 2, must also be 

considered. When prioritizing endpoints for assessment, the ecologists and economists must 

work together to agree to the relative importance of their respective criteria for the overall 

assessment. The economic selection criteria include: 

• Expected magnitude of the change in the economic value of one benefit endpoint relative 

to other endpoints; 

• Anticipated uncertainty associated with the predicted change and value of the change for 

the benefit endpoint relative to other endpoints; 

Page 71 
• Variation in the change to each benefit endpoint under alternative policy scenarios; and 

• Analytical feasibility considerations. 

Analysis Plan 

In an analysis plan, the ecologists describe the data and measures that will be used to evaluate 

the benefit hypotheses, i.e., the relationships between proposed actions and assessment 

endpoints. Measures are identified for exposure, ecosystem and receptor characteristics, and 

effects. Exposure can be quantified or qualitatively characterized (e.g., for chemical 

contaminants, how much enters the environment and how it is distributed, including its possible 

degradation or reaction products). Measures of ecosystem and receptor characteristics identify 

important life history traits (e.g., reproductive cycles, migration patterns, and habitat types) that 

affect the receptors’ potential exposure or the response of assessment endpoints to the change in 

the stressors. Measures of effects quantify the response of the receptors to changes in the 

stressors (e.g., survival, growth, reproduction, and community structure) and help link the effects 

with the assessment endpoints. The analysis plan also specifies how risks will be characterized 

(e.g., qualitative vs. quantitative). 

As indicated earlier, there are two aspects of an assessment endpoint, the entity (e.g., eelgrass) 

and an attribute of that entity (e.g., spatial extent). It is the latter aspect that must be measured 

(quantitatively or qualitatively and directly or indirectly). Spatial extent of eelgrass can be 

quantified by aerial photography; however, for some assessment endpoints, such as songbird 

population (an entity) and abundance (an attribute) as a result of pesticide ingestion, it may be 

difficult to estimate population losses due to mortality if the birds are able to fly away before 

dying. 

Assessment Endpoints and Measures Specified in the Analysis Plan 

An ecological benefits assessment is to be conducted for adding a waste-water treatment 

process and sediment retention ponds at a pulp mill on a river in the Pacific Northwest. 

One assessment endpoint may be Coho salmon breeding success and fry survival. Possible 

measures of the effects of reduced loading of toxic substances to the river on the fish may 

include: egg and juvenile response to low dissolved oxygen, response of adults to change in 

river currents and flow, and adult spawning behavior and egg survival in response to 

sedimentation and contamination. Measures of the ecosystem and receptor (fish) 

characteristics include: water temperature and turbidity, abundance and distribution of 

breeding substrate, food sources for juveniles, variations in abundance, reproductive cycles, 

and laboratory tests for reproduction, growth, and mortality. Measures of exposure could 

include contaminant concentrations in water, sediment, and fish, and dissolved oxygen 

levels in the water. 

As described in Chapter 2, during the prioritization of ecological and economic endpoints, 

ecologists and economists will have discussed what information is required by the economist and 

if that information can be derived from or developed during the ecological assessment. During 

Page 72 
the analysis design phase, ecologists and economists formalize what information is needed by 

economists and determine how and when that information will be provided. They must also 

agree on how changes will be described or measured (e.g., from what baseline, under what 

scenarios, at what level of spatial or temporal detail) and how any limitations or uncertainties 

will be represented. In addition, coordination is required to ensure analytic compatibility 

between the ecological and economic assessments: 

• The baseline from which effects are measured and the specific scenarios or policy options 

to consider must be consistent between the ecological assessment and the economic 

benefit analysis. 

• The analysis plans put in writing the assessment design, the analyses that will be 

conducted, data needs, measures, models to be applied, and statistical techniques to use. 

Both the ecological and economic analysis plans should specify what will be measured 

and how changes in endpoints will be expressed. 

• The spatial area of consideration defined by the ecological benefits assessment serves as 

the starting point for defining the spatial limits of the economic analysis. It generally will 

be true that the spatial scale defined for the ecological analysis will be larger than the 

spatial scale at which the proposed actions can be described. Because the economic 

analysis focuses on human uses associated with ecological resources and humans are 

more mobile than plants and animals, the economic analysis might consider a broader 

spatial area than that defined by the ecological assessment. 

• Both the time horizon and the time step for analysis need to be compatible between the 

ecological and economic analyses. That requirement does not mean that they must be the 

same. For example, the time steps over which economic values of the ecological changes 

are assessed could be shorter than the time steps over which changes in the ecosystem are 

assessed. It is important to account for benefits for future generations in both 

assessments. 

• The economic benefit’s analysis should recognize the uncertainties in the ecological 

assessment process as well as the uncertainties inherent in economic analysis. The level 

of uncertainty in the ecological benefit assessment process is often substantial because 

secondary and tertiary indirect effects are more difficult to estimate than primary or 

direct effects. 

4.3 PHASE II: ANALYSIS PHASE 

Once the analysis plan is complete and the ecologists and economists agree that the analyses and 

measures will be compatible, the actual analyses can begin. In general, the ecological 

assessment must be conducted first to provide the inputs on predicted changes in ecological 

endpoints for the economic assessment. The ecological exposure and response assessments are 

conducted by the ecological risk assessment team independent of the economists. In other 

words, if problem formulation and planning for the analyses are done correctly, there should be 

no need for communication between the economists and the ecologists during the ecological 

Page 73 
analyses. Often, however, unexpected data gaps or unexpected interim modeling results might 

require discussions between the ecologists and economists to resolve such issues. 

The analysis phase consists of the technical evaluation of data to reach conclusions about 

ecological exposure to the stressor and the relationship between the stressor and ecological 

effects (U.S. EPA, 1998). During analysis, risk assessors use measures of exposure, effects, and 

ecosystem and receptor attributes to evaluate questions and issues that were identified in 

problem formulation. 

The analysis phase is composed of two activities: characterization of exposure and 

characterization of ecological effects (U.S. EPA, 1998). These assessments are usually 

conducted simultaneously, and interaction between the scientists conducting them is 

recommended. 

Exposure Assessment Nitrogen Loading in the Chesapeake Bay 

Characterization of exposure in a The Chesapeake Bay is eutrophic, with excess algal risk assessment identifies the 

growth causing declines in fish populations. Several source(s) of the stressor, the spatio­ possible sources of excess nutrients have been temporal distribution of the stressor identified: in the environment, and the contact 

or co-occurrence of the stressor with 

• Atmospheric deposition ecological receptors. Many benefits 

• Run-off from agricultural land assessments will evaluate the 

• Industrial waste streams removal of a stressor from the 

ecosystem; however, similar Although fertilizer runoff is the most obvious source analyses are required to estimate 

which existing exposures might be of the pollution, atmospheric deposition, which may 

originate many miles from the watershed, has been reduced or eliminated by a proposed demonstrated to be a significant loading factor. Any action. The exposure assessment 

should identify the source of the activity proposed to reduce nitrogen loading from

one of these sources should be evaluated in stressor and the complete pathway conjunction with estimates of the loading from the by which it is acting upon the other sources. receptor. A complete pathway 

indicates that a stressor is released 

from a source, is present at a level 

that may cause an effect, and that the receptor is present and susceptible in the ecosystem and cooccurs in time and space with the stressor. Exposure assessment may start with the source when 

it is known, but in cases where the source is unknown, the analysis may attempt to link observed 

contact of the stressor (e.g., a chemical contaminant) with the receptor (e.g., fish) to a source. 

Contaminant residue levels in fish are examples of observed contact. 

In addition to establishing the original or current source of the stressor, the stressor should be 

described in terms of its distribution in time and space. Several factors that may be considered in 

describing a stressor include: 

Page 74 
• Intensity - How much of the stressor is in the environment and at what concentration or 

magnitude? It may be necessary to determine the persistence of the stressor if the 

concentration is not the same at the source as it is at the receptor. 

• Duration - Is the stressor present for a short time or an extended period of time, and how 

is the time defined (hours, days, years)? 

• Frequency - Is the stressor occurring as a single event (chemical spill or volcanic 

eruption), intermittent (pesticide spraying twice a growing season), or continuous ? 

• Timing - What is the occurrence of the stressor relative to biological cycles (e.g., if it 

affects reproduction, is it present during the breeding season or is it present when animals 

are in hibernation)? 

Source of Stressors in Waquoit Bay 

Multiple potential sources were identified for the many stressors acting upon the Bay. Some 

of the sources were local, others were regional. Among the sources of the stressors to the 

Bay are: 

• Cranberry cultivation, which releases nitrogen fertilizers, animal wastes, and 

pesticides; 

• Local and regional atmospheric deposition of nitrogen and toxic contaminants, 

including mercury; 

• Residential development, which results in releases of nutrients from fertilizer and 

septic systems, habitat loss from housing and road construction, and altered 

groundwater flow due to increased impervious surfaces and the number of wells; 

• Industrial discharges to groundwater from a military installation; 

• Sewage treatment facilities and runoff of nutrients and contaminants entering the 

surface waters; and 

• Marine activities that alter habitat, increase contamination, disturb sediments and 

shorelines, dredging, and increased fish and shellfish harvesting. 

Thus to remove many of the stresses on the Bay, actions will be needed in many of the 

community sectors. A screening-level economic assessment of the ecological benefits of 

reducing the stresses might be conducted for the purpose of helping the local governments to 

prioritize activities to reduce the stresses. 

• Location - What is the physical area over which the stressor acts? The stressor may act 

over a very limited area (application of a pesticide in a specific area), or it may act over a 

large area ( tropospheric ozone). What types of habitats are affected (e.g., nesting or 

spawning habitat). 

Page 75 
Many stressors have natural counterparts (e.g., biogenic sources) or multiple sources. The 

characterization of these other sources can be an important component of the analysis. Whether 

alternative sources are analyzed in a given assessment, however, depends on the objectives 

articulated during problem formulation. 

Describe the Distribution of the Stressor in the Environment 

The spatial and temporal distribution of a chemical stressor(s) in the environment is described by 

evaluating the pathways the stressors take from the source to the receptor (e.g., what is the 

medium to which the stressor is released — air, soil, or water — and does it move from one 

medium to another? For example, if a chemical is released to water, does it vaporize?). For 

physical stressors that directly alter or eliminate natural habitats, the temporal and spatial 

distribution of the changed environment should be described (e.g., for how many miles 

downstream from the dredging is turbidity in the water column evident?). For biological 

stressors, the distribution may be more complex. These stressors have the ability to reproduce in 

suitable environments, and do not necessarily rely on passive transport by wind, water, or gravity 

to disperse or move to a suitable habitat. Therefore, when identifying the exposure pathways for 

biological stressors, both active and passive modes of dispersal need to be considered. 

Furthermore, the ability of the biota to reproduce in favorable habitats can alter the relative 

importance of alternate exposure routes. 

Examples of Biotic Interaction 

Metabolism: Several bacteria have been genetically engineered to be particularly useful in 

degrading petroleum. These organisms are able to use petroleum as a food source and break 

down the oil to more environmentally benign compounds. In some cases, metabolism of a 

compound may result in a toxic substance. For example, inorganic mercury compounds 

may be metabolized by microorganisms to methylmercury, which is very toxic. 

Bioaccumulation: Many chemicals that are lipophilic (fat-loving) such as polychlorinated 

biphenyls (PCBs), dioxins, mercury, and cadmium, are readily absorbed and are retained in 

fatty tissues. This way, these chemicals can enter the food chain and affect organisms that 

have been directly exposed. 

The environmental fate of a stressor depends on several factors: 

• Distribution: Once in the environment, where does the stressor go? Stressors may be 

released to or formed from various environmental media. A pollutant released to water 

may partition to the sediment, remain in the water column, or concentrate in the biota. 

Different physical forms of a stressor may partition to different media. 

• Transport: When released or formed, a stressor may be transported from the source. 

Transport occurs via air, water, soil, or biological carrier. Distribution and transport are 

closely related, and are frequently modeled to provide an estimation of where a stressor 

Page 76 
can be found in the environment. The physical, chemical, and biological characteristics 

of both the stressor and the receiving environment determine the transport and 

distribution of the stressor in the environment. 

• Degradation or Transformation: Degradation may occur via biotic processes 

(metabolism), or abiotic processes (transformation by exposure to light or water). 

Degradation implies that a stressor is being physically changed into another simpler 

entity. Transformation may be a gradual or incomplete process (precipitation of a crystal 

from a complex solution). Degradation products and transformation products can also be 

toxic, perhaps more so. For example, elemental mercury released into the environment is 

transformed into methylmercury by microbes in certain aquatic environments. 

Methylmercury is more toxic than elemental mercury, and it is more readily 

bioaccumulated. 

Identifying the distribution, transport, degradation, or transformation processes to which a 

stressor is subject provides an indication to what extent the stressor is likely to act upon a 

potential receptor. It may be possible to show that a stressor is unlikely to affect a receptor given 

its environmental fate and transport. 

The formation and subsequent distribution of 

secondary stressors may be important depending Examples of Secondary Stressors on the objectives of the assessment. For 

chemicals, the evaluation of secondary stressors Chemical: Aldicarb is toxic to mammals usually focuses on metabolites or degradation but not very persistent in the environment. products. Physical disturbance of the However, it is rapidly degraded to aldicarb environment can also lead to secondary sulfone, which is toxic, very persistent, and stressors. Several methods may be used to moves through the soil to the groundwater understand the distribution and environmental where it may remain for years. fate of a stressor and characterize the potential 

exposure of specific receptors to the stressor. Physical: Dredging of a waterway not only Ideally, direct monitoring by collecting and causes loss of habitat for the organisms at analyzing environmental (including biological) the site of the activity, but may result in samples is preferred. Monitoring should be severe turbidity of the water and excessive designed to define the area over which the sedimentation down-current. stressor may be acting and characterize spatial 

and temporal variability in the level of stressor 

(including its degradation products). 

Where monitoring information is lacking or difficult to obtain, models may be used to estimate 

exposure to a stressor. Fate and transport models are commonly used to predict the amount that 

is distributed over a geographic area or the amount of degradation that may be expected over a 

period of time. These models, preferably based on or verified by actual monitoring data, 

generally use the physical, chemical, and biological properties of the stressor as well as the 

environment of concern to characterize the exposure of the stressor to a receptor. This 

characterization should include spatial extent, intensity, frequency, timing, and location of 

exposure. Typically, a combination of monitoring and modeling is used to determine the stressor 

levels. 

Page 77 
Describe the Contact or Co-occurrence with the Receptors 

The exposure assessment must also include an analysis of how the receptors are exposed to the 

stressor (i.e., a pathway by which the stressor acts upon the receptor must be identified). In 

many cases, it is not possible to establish direct causality due to the lack of appropriate 

information. Therefore, it may be necessary to extrapolate or assume that a pathway exists. 

However, if the analyst can demonstrate that a pathway from source to receptor is not plausible, 

then it may be assumed that the receptor will not be affected by the stressor. 

Characterizing the ecosystem on which the stressor is expected to have an impact will assist in 

determining the nature and extent of exposure, and ultimately the adverse effects that may occur. 

If a chemical affects only hardwood trees, but the surrounding area has only softwood trees, any 

observed damage to the softwood trees is unlikely to be the result of the chemical. 

Ecological components may be characterized in many ways, including: habitat, predator/prey or 

feeding relationships, reproductive cycles, and cyclic/seasonal activities. An important 

consideration is at what level of biological organization should an assessment be conducted to 

yield the most useful information? Selection of the best level of organization for an assessment 

must take into account many factors, including tools available for economic analysis if an 

economic analysis is to be conducted. For example, a stressor may cause adverse effects in 

many species in a community, and those effects may be exacerbated or reduced at higher trophic 

levels, depending on the nature of the stressor. A classic example of an ecological stressor that 

is best assessed at the community-level is the bioaccumulation of DDT through the food chain. 

Chapter 3 provides more information on the strengths and limitations of a benefits assessment at 

each biological level of organization. 

It is also important to know the characteristics of the potential receptors. For example: 

• Are they present on a permanent basis (e.g., trees), or are they migratory (e.g., many 

species of birds)? 

• Can and do receptors avoid exposure (i.e., are they capable of detecting the stressor and 

of movement to avoid it)? 

• What are population parameters, such as the size and distribution of the receptors? 

• Is the population particularly vulnerable (e.g., nesting or molting) when exposure is most 

likely to occur? 

• What are the most important physical and temporal parameters (e.g., seasonal and diurnal 

changes in temperature; does the lake freeze in the winter?)? 

Exposure can be described in several different ways, depending on how the stressor causes 

adverse effects: 

Page 78 
• Co-occurrence of the stressor with receptors. Co-occurrence is particularly useful for 

evaluating stressors that can cause effects without actually contacting ecological 

receptors. 

• Contact of a stressor with receptors. Many stressors must contact receptors to cause an 

effect. For example, fish must come in contact with the bacterium Pfiesteria piscicida 

before they become sick or die. 

• Uptake of a stressor into a receptor. Some stressors must not only be contacted, but 

also internally absorbed. For example a chemical that causes liver tumors in fish must 

first be absorbed through the gills to reach the liver to cause the effect. Uptake can vary 

on a situation-specific basis, because it depends on the properties of the stressor (e.g., its 

chemical form), the properties of the receptor (e.g., its physical characteristics and 

health), and the location where contact occurs. 

When the analyses and supporting documentation have been completed, the exposure assessment 

should provide a description of the amount of the stressor that is in the environment, how it is 

able to act on a receptor, and a characterization of the receptor that would or could be affected. 

Exposure assessment is one of the more difficult aspects of an ecological risk or benefits 

assessment, and often introduces the largest uncertainties into the assessment. EPA guidelines 

and other reference materials on conducting exposure assessment, including the use of 

probabilistic methods, should be consulted (e.g., U.S. EPA, 1988, 1989a, 1989b, 1992, 1993a, 

1997b, 1997c; Suter, 1993 ). 

Effects Characterization 

An ecological effects characterization describes the relationship between the stressor 

characteristics (e.g., timing, frequency, magnitude, spatial extent) and the magnitude of the 

resulting ecological effects. The ecological effects characterization indicates the levels of 

exposure that elicit different responses (i.e., the stressor-response relationship). Many stressors 

do not affect all receptors in the same way. In Waquoit Bay, for example, nitrogen loading is a 

significant stressor. Increased nitrogen levels in the Bay result in excessive algal abundance that 

has two effects: (1) shading of eelgrass by the algae, which prevents photosynthesis and kills the 

eelgrass, and (2) decreased oxygen levels in the water that causes physiological stress, 

suffocation, and increased predation on the finfish. In this case, there are several ecological 

effects that can be attributed directly or indirectly to nitrogen loading and that are expected to 

reverse once the nitrogen loading is reduced to more natural levels. The ecological effects 

characterization involves three steps: determining (e.g., quantifying) the stressor-response 

relationship(s), evaluating causality, and linking the measure of effects to assessment endpoints. 

Determine the Stressor-Response Relationship 

Once the receptors and stressors of concern have been defined and plausible exposure scenarios 

have been identified, the next step is to identify those receptors for which stressor- response 

information would be most useful for the ecological effects assessment. Stressor-response 

analysis is often used for chemical stressors such as toxic substances. However, the technique 

Page 79 
can be applied to many stressors and effects, such as increasing levels of microorganisms and 

disease, increasing water temperature and enzyme inactivation, or habitat loss and reproductive 

failure. This type of analysis is particularly valuable, because it describes effects as a function 

of the level of stress. For example, a slight increase in temperature (stressor) in a given stream 

may lead to a significant decline in trout abundance (response), but only a minor decline in algal 

abundance. If the temperature continues to increase, however, algae will also eventually decline 

in abundance. 

Measuring Stressor-Response Relationships 

It is difficult to determine whether algae are alive or dead. However, it is relatively easy to 

measure chlorophyll content both in the laboratory and in the field. Therefore, a change in 

chlorophyll content is often used to measure algal response to stressors, such as increased 

temperature, decreased light, or toxic chemicals. 

Certain types of pesticides are toxic to birds and animals, because they inhibit the enzyme 

cholinesterase, which is necessary for proper neurologic function. It is possible to establish 

a dose-response relationship between the amount of pesticide ingested and the effects of 

cholinesterase inhibition. Relationships may range from changes in blood cholinesterase 

levels with no obvious nerve effects to relatively mild tremors to convulsions and death. 

Stressor-response analysis often provides a quantitative characterization of the stressor and 

effect. For an ecological benefits assessment where the expected magnitude of the change in 

ecological assessment endpoints is needed, full stressor-response curves are needed. In other 

words, a single point on such a curve (e.g., a 50th percentile or 90th percentile response) is not 

useful for benefits assessments. 

Stressor-response relationships are not always linear (e.g., an increase in the stressor will not 

necessarily result in an equal increase in receptor response). For some stressors, a threshold may 

exist below which no response is evident. For example, small increases in water temperature 

may not adversely affect trout – growth may actually be enhanced – but progressively higher 

temperatures will impair growth and, if high enough, result in death. Some stressors may have 

disproportionate ecological effects if the receptors are already subject to another stressor. If deer 

are starving because of deep snow covering their food, the introduction of wolves may reduce 

the deer population by greater numbers than expected. 

Stressor-response data are needed at the biological level of assessment, for example, at the 

population level, the relationship between the stressor and a population-level measure such as 

population density can provide an adequate basis for an assessment of population changes. In 

some cases, stressor-response profiles are estimated from measures at lower levels of biological 

organization (e.g., individual level) based on models (e.g., various population models). For 

some stressors, a quantitative characterization may be difficult to develop. In these cases, a 

qualitative characterization may be used. See Chapter 3 for more discussion on assessments 

conducted at different levels of biological organization. Stressor-response information is 

typically obtained from laboratory or field studies. 

Page 80 
Evaluate Causality 

Without a sound basis for linking cause and 

effect, the uncertainty associated with the 

conclusions of the ecological risk 

assessment is likely to be high. For 

example, many seal populations have 

suffered from epidemics of a distemperlike disease. While several causes 

(stressors) have been suggested and 

studied, including pollution-impaired 

immune systems, warm ocean 

temperatures, reduced food supply, and 

pollution-impaired reproductive systems, 

none have been definitively linked to 

declining seal populations (U.S. EPA, 

1992b). Therefore, while the assessment 

endpoint can be identified for the receptors 

(i.e., a change in seal abundance or 

reproductive success), the potential benefits 

of removing any single stressor cannot be 

quantified. 

The following criteria may be used for 

evaluating causality (U.S. EPA, 1998): 

Criteria strongly affirming causality: 

• Strength of association 

• Predictive performance 

• Demonstration of a stressorresponse relationship 

• Consistency of association 

Identifying Causes for Declines in 

Neotropical Migrant Bird Species 

Populations of neotropical migrant bird 

species appear to be in decline in many areas 

of the United States. These birds, such as the 

Blackburnian warbler, eat insects and live in 

the interior of large blocks of forest where 

they breed. They migrate south in the winter, 

following their food supply. 

The risk hypothesis is that population decline 

is caused by forest fragmentation in North 

America and deforestation in tropical South 

America. Forest fragmentation results in loss 

of core forest areas and creation of additional 

forest edge habitat which in which the 

breeding success of the birds is lower due to 

predation and parasitism by species adapted to 

open and edge habitats. 

Data (taken from previous studies) were 

gathered to assess the susceptibility of 

neotropical migrant species to edge effects, 

island effects, and the loss of wintering habitat 

in the tropics. Further monitoring was 

recommended, including the development of 

databases to collect additional data on these 

birds. 

Page 81 
Criteria providing a basis for rejecting causality: 

• Inconsistency in association 

• Temporal incompatibility 

• Factual implausibility 

Other relevant criteria: 

• Specificity of association 

• Theoretical and biological plausibility 

Link the Measures of the Effects to the Assessment Endpoints 

Assessment endpoints express the environmental values of concern for a risk assessment, but 

cannot always be measured directly. When the measures of effect differ from assessment 

endpoints, sound and explicit linkages between the two are needed. 

The following are examples of extrapolations that risk assessors may use to link measures of 

effect to assessment endpoints (U.S. EPA, 1998): 

• Between similar organisms (e.g., bluegill to rainbow trout); 

• Between responses (e.g., mortality to growth or reproduction); 

• Between different sources of data (e.g., laboratory to field data); 

• Between geographic areas (e.g., northeastern U.S. to northwestern U.S.); 

• Between spatial scales (e.g., stream to river); and 

• Between temporal scales (e.g., data for short-term effects to longer-term effects). 

During the development of the analysis plan in the problem formulation phase (Phase I), the 

ecological assessment team identified the extrapolations that would be required between 

assessment endpoints and measures of effect. Decisions about specific extrapolations are usually 

based on the scope and nature of the risk assessment, resources available for conducting the 

assessment, and the amount of uncertainty that is acceptable. During the analysis phase, the 

assessors implement these extrapolations. However, they should reconsider all available data to 

determine whether the plan should be modified. For example, the exposure characterization may 

indicate different spatial or temporal scales than originally anticipated. If a stressor persists for 

an extended time in the environment, it may be necessary to extrapolate short-term responses 

over a longer exposure period and population-level effects may become more important. 

The goal of the analysis phase is to provide sufficient information such that it is possible to 

characterize the changes in ecological assessment endpoints specified during Problem 

Formulation. 

Page 82 
Characterizing Uncertainty 

Uncertainty evaluation is an ongoing issue 

throughout the analysis phase. The purpose 

of an uncertainty analysis is to formally 

recognize that the ecological risk assessment 

is constructed upon incomplete knowledge 

and to explain the implications. Specifically, 

the uncertainty analysis characterizes both the 

qualitative and quantitative uncertainties 

associated with the input values and carries 

those uncertainties through to the estimated 

exposure and ecological effects. 

Any uncertainty analysis need not always be 

expressed mathematically. Instead, a 

qualitative description may be used, such as 

indicating that the animal tested may not be 

the best surrogate for animals actually 

exposed to a stressor. This frequently occurs 

in wildlife toxicity testing where the 

laboratory animal may be more or less 

sensitive than other species in the wild. 

Each of the extrapolations listed in the 

previous subsection also introduces 

Uncertainty Factors 

Uncertainty factors may be quantitative or 

qualitative depending on their application. 

In the development of a conceptual model 

for the benefits assessment, there may be 

uncertainty associated with the assumptions 

used for the model. Examples may be the 

use of a well characterized species as a 

surrogate for a species that is less well 

characterized (e.g., use of coyotes rather 

than wolves). A pathway may not be 

clearly defined from the source of the 

stressor to the receptor. For example, a 

species of bird exhibit impaired 

reproduction. The initial risk hypothesis 

was that loss of habitat from timber cutting 

was responsible for the impairment. 

Alternative hypotheses (e.g., the birds are 

exhibiting reproductive effects as a result of 

runoff from the timber cutting exposing 

contaminated soil) also should be 

considered. 

uncertainty. Other sources of uncertainty in an ecological risk/benefit assessment include, but 

are not limited to: 

• Sampling variability; 

• Inability to obtain appropriate samples (this may be of concern if the organism is 

endangered or difficult to identify or collect); 

• Lack of knowledge about combined effects of multiple stressors; and 

• Nonlinear behavior of complex systems. 

Quantitative measures of uncertainty are often difficult (and sometimes impossible) to provide; 

when this is the case, the assessors should try to characterize uncertainty in a qualitative manner 

as completely as possible. This ensures that economists, policy makers, and others who use the 

results of the ecological risk assessment have a sense of the assessment’s strengths and 

weaknesses. 

Methods for analyzing and describing uncertainty associated with an ecological risk assessment 

range from simple to complex and are beyond the scope of this document. For further reading 

Page 83 
on the topic of uncertainty analysis, see EPA publications (1992, 1997a, 1997b) and discussions 

by Suter (1993) and Suter et al. (2000). 

4.4 PHASE III: RISK/BENEFIT CHARACTERIZATION 

The final step in an ecological benefits 

Stressor-Response Relationships assessment is estimation and characterization 

of the changes in the ecological assessment Examination of a time series of indicators of endpoints specified by the conceptual model the health of the forests in the northeastern and the analysis plan. In this step, the United States suggested that gypsy moths characterization of exposure and might be playing a significant role in the characterization of ecological effects (Phase observed decline in forest condition. To III - Analysis) are integrated to provide an estimate the costs and benefits of indication of the changes in the ecological controlling gypsy moths, information on the assessment endpoints and associated relationship between gypsy moth abundance uncertainty. For a benefits assessment, single and forest condition was sought. The point estimates of risk (e.g., threshold for literature revealed that a small number of effect) is not useful; the hazard quotient gypsy moth larvae may cause minor damage approach is not applicable. Instead, the risk to the foliage on some trees. However, a and benefit estimation should be based on the 

larger infestation can result in stunted tree entire stressor-response relationship and the growth or even tree death if the larvae eat probability of exposure, and decided based on enough leaves where trees cannot sustain the process models. In other words, the their photosynthetic requirements. The ecological risk assessment team estimates the density of gypsy moths can be directly likely degree of change in the assessment related to tree damage, up to and including endpoint from the probability estimates of death. exposure-probability function and from the 

stressor-response curves. 

Monte Carlo simulations or other probabilistic approaches for incorporating variability and 

uncertainty can be used to estimate the probability of exposure at various levels. The Team also 

should attempt to provide uncertainty bounds on that estimate and the likelihood that actual 

responses would be greater or less than those predicted. This determination can be qualitative or 

quantitative. For further discussion of probabilistic assessments, see other EPA documents 

(1997a, 1997b). 

Process models are mathematical expressions that represent our understanding of the 

mechanistic operation of a system under evaluation. A major advantage of using process models 

for a benefits assessment is the ability to consider “what if” scenarios, and to forecast beyond the 

limits of the observed data that constrain risk estimation techniques based on empirical data. For 

example, process models may be used to extrapolate from individual-level to population- and 

ecosystem-level effects. These models may also be of use in estimating indirect effects on the 

assessment endpoints and the probable rate of recovery. A variety of process models are 

available for both terrestrial and aquatic ecosystems (e.g., RAMAS, Aquatox). Because process 

models are only as good as their assumptions, they should be treated as hypothetical 

representations of reality until appropriately tested with empirical data. 

Page 84 
After the magnitude and likelihood of changes in the ecological assessment endpoints have been 

estimated, the results are ready to supply to the economic analysts. The ecological assessment 

team should characterize both the beneficial and adverse effects that could accrue from an action 

so that the economists can estimate “net” benefits. The ecological assessment team needs to 

provide the economic analysts with full descriptions of the potential range in natural variability 

that might be expected in the assessment endpoint and the uncertainties associated with their 

estimates of changes to the assessment endpoints. Where the analysis plan called for 

quantitative assessments of variability and uncertainty for specific assessment endpoints, the 

ecological assessment team should provide quantitative estimates to the extent possible. For all 

other assessment endpoints, the ecologists and economists agreed to pursue only a qualitative 

assessment of variability and uncertainty. 

There are several other parameters that must be addressed to characterize changes in the 

ecological assessment endpoints with sufficient specificity to be used in the economic 

assessment. The magnitude of effect needs to be specified in terms of geographic coverage, the 

degree of change per unit area, the probability of changes of that degree or higher (or lower), and 

the time-frame over which the change would be expected to occur. Again, quantitative or 

qualitative error bounds should be provided for these parameter estimates. 

This document has focused on the economic assessment of ecological benefits. The ecological 

assessment team might have included assessment endpoints that were not going to serve as 

inputs to the economic analysis. The team would prepare and present these results to the 

audience for which they were intended. 

The next chapter discusses how economists define the value of changes in an ecological 

assessment endpoint, and subsequent chapters discuss the steps associated with economic 

benefits analysis and the specific methods for the economic valuation of ecological benefits. 

References and Further Reading 

Bartell, S.M.; Gardner, R.H.; and O’Neill, R.V. 1992. Ecological Risk Estimation. Chelsea, 

MI: Lewis Publishers. 

Norse, E. 1990. Threats to Biological Diversity in the United States. Report prepared for the 

U.S. EPA, Washington, DC, by Industrial Economics, Contract No. 68-W8-0038, Work 

Assignment 115. 

Noss, R.P.; LaRoe, E.T.; ans Scott, J.M. 1995. Endangered Ecosystems of the United States: A 

Preliminary Assessment of Loss and Degradation. Washington, DC: U.S. Department of the 

Interior, National Biological Service. 

Principe, P.P. 1995. Ecological benefits assessment: A policy-oriented alternative to regional 

ecological risk assessment. Human and Ecological Risk Assessment 1(4):423-435. 

Suter, G.W. II. 1989. “Ecological Endpoints.” in Warren-Hicks, W., B.R. Parkhurst, and S.S. 

Baker, Jr., eds. Ecological Assessment of Hazardous Waste Sites: A Field and Laboratory 

Page 85 
Reference Document. EPA/600/3-89/013. Corvallis Environmental Research Laboratory, 

Oregon. 

Suter, G.W. II. 1993. Ecological Risk Assessment. Boca Raton, FL: Lewis Publishers. 

Suter, G. W., Efroymson, R.A., Sample, B.E., Jones, D.S. 2000. Ecological Risk Assessment for 

Contaminated Sites. Boca Raton, FL: Lewis Publishers. 

U.S. EPA. 1988. Superfund Exposure Assessment Manual. Washington, DC: Office of Solid 

Waste and Emergency Response Directive 9285.5-1; EPA/540/1-88/001. 

U.S. EPA. 1989. Superfund Exposure Assessment Manual -- Technical Appendix: Exposure 

Analysis of Ecological Receptors. Athens, GA: Office of Research and Development, 

Environmental Research Laboratory (December). EPA/600/3-88/029. 

U.S. EPA. 1990a. Ecosystem Services and Their Valuation. Prepared by RCG/Hagler, Bailly, 

Inc., for the Office of Policy, Planning, and Evaluation. Washington, DC: U.S. EPA. 

U.S. EPA. 1990b. Biological Criteria: National Program Guidance for Surface Waters. 

Washington, DC: U.S. EPA. EPA/440/5-90/004. 

U.S. EPA. 1991. The Watershed Protection Approach Framework Document. Office of 

Wetlands, Oceans, and Watersheds. Washington, DC: U.S. EPA. 

U.S. EPA. 1992. Guidelines for Exposure Assessment. Federal Register. 57: 22888-22938 

(May 29). 

U.S. EPA. 1993a. Wildlife Exposure Factors Handbook Volumes I and II. Washington, DC: 

Office of Research and Development; EPA/600/R-93/187ab. 

U.S. EPA. 1993. Guidance for Specifying Management Measures for Sources of Nonpoint 

Pollution in Coastal Waters. Office of Water. Washington, DC: U.S. EPA. EPA 840/B-92/002. 

U.S. EPA. 1993. A Guidebook to Comparing Risks and Setting Environmental Priorities. 

Washington, DC: U.S. EPA. EPA/230/B-98/003. 

U.S. EPA. 1994a. Managing Ecological Risks at EPA: Issues and Recommendations for 

Progress. Prepared by M.E. Troyer and M.S. Brody. Washington, DC: U.S. EPA. EPA/600/R94/183. 

U.S. EPA. 1994b. Toward a Place-Driven Approach: The Edgewater Consensus on an EPA 

Strategy for Ecosystem Protection. Ecosystem Protection Workgroup. Washington, DC: U.S. 

EPA. March 15 Draft. 

U.S. EPA. 1996. Waquoit Bay Watershed Ecological Risk Assessment Problem Formulation. 

U.S. EPA. Review Draft. 

Page 86 
U.S. EPA. 1997a. Priorities for Ecological Protection: An Initial List and Discussion 

Document for EPA. Washington, DC: U.S. EPA. EPA/600/S-97/002. 

U.S. EPA. 1997b. Policy for Use of Probabilistic Analysis in Risk Analysis. Office of the 

Administrator, Washington, DC. May. 

U.S. EPA. 1997c. Guiding Principles for Monte Carlo Analysis. Risk Assessment Forum, 

Washington, DC. EPA/630/R-97/001. March. 

U.S. EPA. 2000. Assessing the Neglected Ecological Benefits of Watershed Management 

Practices: A Resource Book. Assessment and Watershed Protection Division, Office of Water. 

Washington, DC: U.S. EPA. 

Page 87 
5.0 BACKGROUND THEORY ON VALUING CHANGES TO 

ECOLOGICAL RESOURCES 

This chapter discusses how economists define the value of an ecological resource and then 

estimate the value of a change to such a resource. This chapter provides a basic introduction to 

welfare economics as applied to the valuation of environmental changes -- it is not, however, a 

comprehensive review of welfare economics. References for further reading on the subjects 

discussed below are provided at the chapter’s end. 

Subsequent chapters will discuss the steps associated with economic benefits analysis and the 

specific methods for estimating benefits, applying the general concepts introduced in this 

chapter. Readers should be aware that subsequent discussions presume an understanding of 

basic economic concepts (e.g., supply, demand). 

5.1 WELFARE ECONOMICS AND THE VALUE OF AN ECOLOGICAL CHANGE 

Welfare economics provides the theoretical basis for estimating the economic benefits of an 

action. Welfare economists assess the value of an action based on its effect on the well-being, or 

level of welfare, of humans. Economic value is based on what people want -- that is, their 

preferences -- and is measured by examining how people trade-off different goods and services. 

Economic theory is based on the assumption that the decisions people make regarding what they 

choose to consume and what they choose to do for activities reflect the values they hold for the 

various goods and services available. 

The anthropocentric perspective of welfare economics implies that the economic value of an 

ecological resource depends on the value humans derive from the resource. Some people take 

affront to this basic premise of welfare economics, arguing that decisions should be based on the 

broader values of community, the impacts to future generations, or the inherent rights of natural 

resources (see for example, Sagoff, 1988). Nonetheless, welfare economic analysis provides 

useful information for making decisions and presently serves as the basis for most 

economic/policy analyses. For this reason, this document focuses on describing the basic 

concepts and techniques of welfare economists. Decisionmakers must, however, keep in mind 

that a welfare analysis is just one approach among many for evaluating a change and should 

consider other perspectives and information when making policy choices. 

The type, quantity, and quality of goods and services available to an individual determine the 

individual’s level of well-being, or level of welfare. Some goods and services are produced by 

industry and purchased by individuals in markets, some are produced within the household, some 

are provided by government, and some are provided by nature or ecological resources. (See 

Chapter 6 for a discussion of the different types of goods and services provided by ecological 

resources). 

The condition of an ecological resource determines the type, quantity, and quality of goods and 

services provided by that resource. As a result, any action that affects an ecological resource, 

such as an environmental regulation or natural resource management activities, will likely also 

Page 88 
affect the goods and services the resource provides, and subsequently the level of welfare of the 

individuals who enjoy those goods and services. Typically, economists estimate the welfare 

change associated with a policy or action using an “effect-by-effect” approach (U.S. EPA, 2000). 

Under this approach, economists measure the change in welfare associated with the change to 

each good and service provided by the ecological resource and sum these individual measures to 

estimate total benefits. The next section describes how economists measure changes in welfare. 

5.2 MEASURING THE BENEFITS OF IMPROVEMENTS TO ECOLOGICAL 

RESOURCES – THE CONCEPT OF WILLINGNESS-TO-PAY 

The economic value of a good is determined by the maximum amount of something else (usually 

money) that an individual is willing to pay to obtain the good. This measure of economic value 

is called “willingness-to-pay” (WTP). For an environmental improvement, WTP is the amount 

an individual is willing to pay to obtain the improvement. An alternative measure, “willingnessto-accept” (WTA) is defined as the minimum amount of money an individual is willing to 

receive in compensation to forgo a benefit, such as an environmental improvement, they would 

otherwise receive. 

The choice between using WTP or WTA to value changes in environmental quality implies 

different assumptions regarding the property rights of individuals experiencing the change. 

Using WTP implies that polluting entities have a right to pollute, so the public must pay them 

not to pollute. Using WTA implies that the public has a right to a clean environment and must 

be compensated for pollution. There also can be significant differences in the estimated value of 

a change measured in terms of WTP or WTA. One reason for this difference is that for an 

environmental improvement, WTP is based on an individual’s level of welfare without the 

improvement, while WTA is based on the level of welfare achieved with the improvement. (See 

Hanley, Shogren, and White, 1997 for further discussion of this issue). 

Additionally, measuring economic value in terms of WTP also does not allow for the possibility 

that certain goods may be "incommensurable" for some individuals, because their WTP is 

constrained by their income level. This constraint of welfare economics imposes an ethical 

assumption that people will always be willing to substitute other goods for ecological resources. 

Although WTA is typically the theoretically correct measure for estimating the benefits of 

environmental improvements, WTP is more commonly used in practice because it is easier to 

measure and estimate (U.S. EPA, 2000). 

For consistency with how goods and services are traded through markets and comparability with 

the estimated dollar costs of an action, economists measure the benefits of an action, such as a 

regulation, in dollar terms using WTP. WTP values reflect individual’s preferences for 

exchanging goods and services. Because preferences are likely to vary from one individual to 

another, WTP values for a change to a particular good or service will vary from one individual to 

another. The total social value of an improvement in a good or service is the sum of the WTP 

across all individuals. 

Although economists are most often asked to value the change in social welfare (measured by 

WTP) associated with a change in a particular good or service provided by an ecological 

resource, they will also sometimes be asked to value the availability or existence of the 

Page 89 
ecological resource itself. For example, an economist might be asked to determine the change in 

social welfare associated with the complete loss of a wetland and all the goods and services it 

provides. In this circumstance, the economist will likely have to value the individual goods and 

services lost separately and sum these benefit estimates. Alternative approaches that attempt to 

estimate the total value of such a resource based on replacement cost or embodied energy (e.g., 

Costanza et al., 1997; Ehrlich and Ehrlich, 1997; Pearce, 1998; Pimentel et al., 1997) have been 

discussed recently, but are not appropriate for an economic benefit analysis because the methods 

are not grounded in economic theory (U.S. EPA, 2000). 

5.3 HOW ECONOMIC BENEFITS OF IMPROVEMENTS TO ECOLOGICAL 

RESOURCES ARE REALIZED 

The economic benefits of an action that affects an ecological resource depends on how the state 

of the ecological resource influences the supply or consumption of the goods and services 

provided or supported by that resource. In evaluating the economic benefits of an action 

affecting an ecological resource, economists consider two possible relationships between the 

resource and the goods and services enjoyed by society: 

• The ecological resource is an input to the production of a good or service, such that the 

state of the ecological resource directly affects the production (or supply) of the good or 

service; or 

• The state of the ecological resource is a characteristic of the good or service, such that the 

state of the ecological resource directly or indirectly affects the demand for (or value of) 

the good or service. 

Ecological Resource as Input to Production 

When an ecological resource serves as an input in the production of another good or service, 

changes affecting the quality or state of the ecological resource can have direct impacts on the 

production or supply of the good or service. “Production” might consist of a natural or 

bioeconomic process, or a man-made or industrial process. For example, a change in the ozone 

concentration in the air will affect the growth rate of plants and, thus, the productivity of 

agricultural crops. Alternatively, an improvement in the water quality of a river that provides 

water used for paper production may reduce the processing costs and, thus, increase the 

productivity of the paper mill. That same change in the water quality of a river might also affect 

the non-market recreational opportunities provided, or “produced,” by the river. The effect of 

the change in the productive process may be realized through changes in the flow of a nonmarket good or service, a change in the price of a marketed good or service, or a change in the 

wage rates or earnings of workers in the affected sector. 

The State of the Ecological Resource as Characteristic of a Good or Service 

When the state of an ecological resource is a characteristic of a good or service, a change in the 

state of the ecological resource affects the demand for that good or service. For example, the 

demand for recreational fishing days at a particular lake is likely to change if the water quality of 

the lake is improved. Similarly, the demand for hiking days or scenic views may increase as a 

Page 90 
result of an improvement in air quality that increases visibility. The change in demand may be 

realized through increased number of visitation days, increased number of users, or increased 

spending to make use of a good or service provided by the ecological resource. 

5.4 ESTIMATING WILLINGNESS-TO-PAY 

There are several techniques used to estimate WTP for changes to goods and services provided 

by ecological resources. The technique employed depends on the type of good or service 

affected. The techniques used by economists may: 

• Estimate demand and supply curves for the good or service in question; 

• Estimate demand and supply curves for a related good or service; or 

• Estimate WTP based on other observations. 

Market Goods and Services 

The change in social welfare for a given change in the supply or price of a good and service that 

is sold through a market is often approximated by the sum of predicted change in consumer and 

producer surplus. Consumer and producer surplus is represented as the area above the supply 

curve and below the demand curve. These surplus measures are standard and widely accepted 

terms of applied welfare economics. Consumer and producer surplus is derived from market 

data on how much of the good is demanded and produced in the aggregate at various price levels 

and can be easier to estimate than individual WTP. 

Although surplus measures do not, in general, provide a theoretically correct estimate of the 

change in social welfare, they can provide a reasonably accurate estimate of social WTP for 

relatively small price changes (Willig, 1976). The estimate is less reliable, however, for changes 

in the quality or quantity of goods and services. Nonetheless, measures of changes in consumer 

and producer surplus are often used as indicators of the economic magnitude of impacts when 

more precise measures are not feasible or practical. 

Non-Market Goods and Services 

The lack of markets and prices for many of the goods and services provided or supported by 

ecological resources often makes it more difficult to estimate WTP for changes to these goods 

and services. For goods and services that are not traded through markets, economists measure 

changes in economic welfare, or WTP, based on changes in human behavior and the decisions 

people make under different circumstances. Economic techniques for non-market goods and 

services estimate individual WTP using either market information for related goods and services 

(revealed preference methods) or direct statements of people’s preferences (stated preference 

methods). Individual WTP estimates are generally averaged and multiplied by the total number 

of affected individuals. 

Chapter 6 provides more information on specific methods for estimating WTP for market and 

non-market goods and services. 

Page 91 
References and Further Reading 

Braden, J.B. and C.D. Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. 

North-Holland, Amsterdam: Elsevier Science Publishers. 

Costanza, R. et al. 1997. “The Value of the World’s Ecosystem Services and Natural Capital.” 

Nature 387: 253-260. May. 

Ehrlich, P. and A. Ehrlich. 1997. Betrayal of Science and Reason. Island Press, Washington, 

D.C. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Hanemann, W.M. 1991. “Willingness to Pay and Willingness to Accept: How Much Can They 

Differ?” American Economic Review. 81(3): 635-647. 

Hanley, N., J.F. Shogren, and B.White. 1997. Environmental Economics in Theory and 

Practice. Oxford University Press, New York, New York. p. 362-364, 395-396. 

Just, R.E., D.L. Hueth, and A. Schmitz. 1982. Applied Welfare Economics and Public Policy. 

Englewood Cliffs, New Jersey: Prentice-Hall, 

Pearce, D. 1998. “Auditing the Earth.” Environment. 40(2): 23-28. 

Pimentel, D., C. Wilson, C. McCullum, R. Huang, P. Dwen, J. Flack, Q. Tran, T. Saltman, B. 

Cliff. 1997. “Economic and Environmental Benefits of Biodiversity.” BioScience. 47(11): 

747-757. 

U.S. EPA. 2000. Guidelines for Preparing Economic Analyses. U.S. EPA, Office of the 

Administrator. EPA/24/R-00/003. September. 

Willig, R. 1976. “Consumer Surplus Without Apology.” American Economic Review 66(4): 

589-597. 

Page 92 
6.0 ECONOMIC ASSESSMENT OF ECOLOGICAL 

BENEFITS 

The first section of this chapter discusses the various components or steps of an economic 

assessment of ecological benefits (“economic benefit analysis”). The steps of an economic 

benefit analysis described here are based on EPA’s (2000) Guidelines for Preparing Economic 

Analyses. Later sections of this chapter provide additional information on the types of economic 

benefit endpoints that might be considered by an economic benefit analysis and the different 

approaches available for estimating the economic value of those benefits. 

6.1 COMPONENTS OF AN ECONOMIC ASSESSMENT OF ECOLOGICAL BENEFITS 

As discussed in Chapter 2, there is an extensive planning phase that precedes the ecological risk 

assessment and the economic benefit analysis. The decisions made during that planning process 

regarding the nature of the decision under study and other criteria guide the economic analysis. 

Of particular importance in designing and conducting the economic benefit analysis is a good 

understanding of the type of information needed by decisionmakers and the nature of the 

information that will be made available from the ecological risk assessment. 

Following the planning phase, the economic benefit analysis begins. The economic benefit 

analysis generally follows an “effect-by-effect” approach (U.S. EPA, 2000). As noted in 

Chapter 2, under this approach economists estimate the benefits associated with the major effects 

of a policy or action separately and then sum together the value estimates for the individual 

effects to arrive at an estimate of the total benefits. 

An economic benefit analysis can be broken down into four steps: 

• Identify and prioritize economic benefit endpoints 

• Describe and quantify effects of the policy or action on the economic benefit endpoints 

• Estimate the value of those effects 

• Summarize and present the results 

The key elements of each of these steps are discussed below. Chapter 2 discusses in detail the 

opportunities for improving the economic benefit analysis by coordinating with the ecological 

risk assessment team in each of these steps. Chapter 2 also discusses the variety of issues that 

must be coordinated with the ecological risk assessment in designing an economic benefit 

analysis: establishing the baseline from which changes are measured, measuring changes to the 

economic endpoints based on the changes to the ecological endpoints, determining the 

appropriate spatial and temporal scale for the analysis, and determining how uncertainty will be 

treated. These discussions are not repeated here. 

Page 93 
6.1.1 Identify and Prioritize Economic Benefit Endpoints 

The process of identifying and prioritizing the economic benefit endpoints affected by a policy 

or action is described only briefly here. A more detailed discussion of this step is provided in 

Chapter 2. 

The identification and prioritization of relevant economic benefit endpoints involves: 

• Developing a list of preliminary economic endpoints; 

• Linking changes to ecological resources to changes in the economic benefit endpoints; 

• Prioritizing the economic benefit endpoints for consideration by the benefit analysis; and 

• Determining how the economic analysis will assess the changes for high priority 

economic benefit endpoints. 

Economists identify potential economic benefit endpoints by understanding the policy or action 

under study, reviewing analyses of similar policies or actions, and working with the ecological 

risk assessment team to understand what ecological changes are expected.Linking ecological 

changes to changes to specific economic benefit endpoints involves extending the conceptual 

model developed by the ecological risk assessment team to include the economic benefit 

endpoints that are expected to be affected. As noted in Chapter 2, the ecological risk assessment 

team can provide valuable assistance to the economists in determining how various economic 

benefit endpoints might be affected. 

Time and resource constraints generally require that the economic benefit analysis focus on fully 

assessing and valuing changes to a limited number of endpoints. Each endpoint’s consideration 

priority is based on decisionmakers’ needs, the expected magnitude of the predicted change to 

that endpoint, the uncertainty associated with the predicted change and anticipated value of the 

change, and the variability of the change to each benefit endpoint under different policy options. 

This prioritization method ensures that important changes that typically are not amenable to a 

monetary assessment are given equal attention by the benefit analysis. For the high priority 

economic benefit endpoints, the economic analyst must determine how the changes to each 

endpoint will be analyzed -- with a qualitative, quantitative, or monetized assessment of the 

change. The choice of assessment method depends upon the relative need for a dollar value 

estimate of benefits, the availability of the necessary data and appropriate quantification and/or 

economic valuation techniques, and the time and resource constraints of the economic benefit 

analysis. (See Chapter 2 for an extensive discussion of these steps.) 

Page 94 
6.1.2 Describe and Quantify Changes to the Economic Benefit Endpoints 

Using results from the ecological risk assessment and information from other data sources, the 

economist describes the changes to the ecological services and values affected (i.e., the 

economic benefit endpoints) and provides information on the magnitude of the changes. The 

ecological risk assessment is responsible for estimating the likely changes to the ecological 

resources. If care has been taken to coordinate during problem formulation (see discussion in 

Chapter 2), the information provided by the risk assessment will be compatible with the needs of 

the economic benefit analysis. 

In addition to the information from the risk assessment, economists may collect additional 

economic data to describe and estimate the effects and economic value of changes to the 

economic benefit endpoints. For example, in describing the impact of a change to an economic 

benefit endpoint, the economist might consider the estimated number of users of the goods and 

services provided by the resource (e.g., number of fishermen, number of visitors, local 

population, national population), the quantity of the good or service provided or used (e.g., 

timber production, commercial fish landings), or some measure of the magnitude of the 

ecological resource itself (e.g., acres, productivity). 

A thorough qualitative, and when possible quantitative, discussion of the changes to the 

economic benefit endpoints is an essential component of the benefit analysis. For endpoints for 

which a monetized assessment is not possible, the qualitative or quantitative assessment provides 

a measure of the good or service’s importance and the degree of change experienced under the 

policy or action. For those endpoints for which a monetized assessment is conducted, the 

qualitative and quantitative discussion of the change that is valued supports the dollar value 

generated by the benefit analysis. Some ecological improvements may also result in economic 

losses. The net of positive and negative economic changes must be calculated in determining the 

benefits of any action. 

6.1.3 Estimate the Value of the Changes 

There are a wide variety of techniques available to estimate the value or change in value of 

specific attributes or services provided by ecological resources. In this step, the economist 

selects the approach that is most appropriate given the attribute or service being analyzed, the 

data available regarding the production or demand for the attribute or service, and the time and 

resource constraints of the study. The most common approach used by EPA analyses is benefit 

transfer, in which value estimates from one study are applied to another situation. 

Regardless of the technique used to estimate the value of changes to the economic endpoints, the 

benefit analysis must describe the source of the value estimate and the degree of confidence in 

the estimate. This is particularly important when using benefits transfer because transferring a 

value estimate to a new situation can only increase the uncertainty associated with the estimate. 

Several different techniques may be used to estimate the benefits associated with changes to 

multiple endpoints. Although using multiple methods may provide more information on the 

value of changes experienced by the economic endpoints, care must be taken to avoid doublecounting of benefits. A careful understanding of the relationships among the various endpoints 

Page 95 
considered is important for identifying potentially overlapping benefits that could lead to doublecounting. 

Many economic valuation techniques estimate the value or benefits associated with a change for 

an individual person (i.e., individual willingness-to-pay (WTP)). To develop an aggregate 

estimate of the social benefits of an action, economists sum individual WTP for the action across 

all affected individuals. For many actions that affect ecological resources, some individuals will 

benefit while other individuals may experience a decline in their individual welfare. For 

example, while removing a dam might improve opportunities for kayaking and other whitewater 

activities, it might also result in the loss of a boating area behind the dam for water skiing and 

fishing. In this type of situation, the benefit of the action is the net total of all gains minus all 

losses in individual welfare experienced by members of society. 

In aggregating benefits estimates across all affected individuals, each individual’s WTP is given 

the same weight in the summation. As discussed in EPA’s Guidelines for Preparing Economic 

Analyses (U.S. EPA, 2000), an equity assessment and impact analysis may be conducted to 

assess the impact of an action on any populations of concern. 

Typically, the economic benefit analysis will examine changes that occur over an extended 

period of time (i.e., longer than a single year). The economic benefit analysis, therefore, may 

need to describe when these changes occur over the time period considered by the analysis as 

well as summarize the value of the change over the whole time period. The standard approach 

for summarizing the value of changes that occur over an extended time period is to sum the value 

of all changes over the time period using discounting. Discounting is commonly used to express 

future costs or benefits in present monetary value. The use of discounting and the choice of an 

appropriate discount rate are complex and highly debated issues that are beyond the scope of this 

document. Of particular concern is the potential effect of discounting ecological benefits on 

resource conservation and intergenerational equity issues. References for additional information 

on discounting are provided in Chapter 7. 

6.1.4 Summarize and Present the Results 

As discussed in Chapter 2, the results of the economic benefit analysis will include the 

prioritized list of economic effects, discuss the criteria used to select the economic endpoints 

examined in detail by the benefit analysis, and discuss how the economic value of the effects was 

assessed. The monetary benefits estimated for some of the changes will be accompanied by the 

qualitative assessment of other benefits that were not monetized. If possible, the qualitative 

assessment should discuss the potential magnitude of the economic benefits for any priority 

endpoints that are not accounted for by the quantitative and monetized assessment. The final 

report should also discuss to some degree the other effects identified that were deemed less 

important to the economic analysis. 

Finally, the results of the economic analysis must disclose any source of error in the analysis and 

the potential impact of such error on the results. The presentation of results should identify any 

possibility of double-counting of benefits, any limitations of the analysis, and any potential 

imprecision and uncertainty associated with the benefit estimates. The analyst should note 

uncertainties in the ecological assessment that are relevant to the economic assessment and 

Page 96 
discuss how these uncertainties may be compounded. In discussing the potential impact of any 

source of imprecision or uncertainty, economists should discuss whether the analysis is likely to 

over- or under-estimate the economic value of benefits. 

The results presented also may include information and details that are needed for other 

analyses. For example, an equity analysis may require information on the geographic 

distribution of effects, the distribution of ecological effects and economic benefits across 

different ethnic or economic classes of the human population, or the distribution of ecological 

effects and economic benefits over time. 

Page 97 
6.2 IDENTIFYING THE SERVICE FLOWS AND OTHER VALUES PROVIDED 

BY AN ECOLOGICAL RESOURCE 

There are numerous types of goods and services provided by ecological resources that have 

economic value to some or all individuals in society (see Chapter 5 for a discussion on defining 

the economic value of 

ecological resources). This 

section discusses the various 

types of goods and services Exhibit 16 

and offers their taxonomy, Proposed Taxonomy of Goods and Services 

which may be useful in Provided by Ecological Resources developing a comprehensive 

list of specific economic Good or Service benefit endpoints. The 

proposed taxonomy for 

generally characterizing the 

goods and services provided Non-Use Value Use Value 

by ecological resources is 

presented in Exhibit 16. 

Some of goods and services 

provided by ecological 

resources are obvious 

Direct, Market Direct, Non-Market Indirect, Non-Market 

Direct Use Indirect Use 

because they are directly 

used or enjoyed by society, Good or Service Good or Service Good or Service 

such as the fish provided by 

a fishery, the timber/lumber 

provided by a forest, or the 

Page 98 
swimming and boating opportunities provided by a coastal area. These types of goods and 

services are defined as direct, market uses, when the good or service is bought and sold through 

open markets, and direct, non-market uses, when the good or service is not bought and sold 

through a market. 

The direct, market uses of an ecological resource are typically the most obvious and most easily 

valued goods provided by an ecological resource because price and quantity information for each 

good and service is generally available. The direct, non-market uses of an ecological resource 

may be readily apparent, such as recreational opportunities, although more difficult to value. 

Valuation of changes to direct, non-market uses is more difficult because the goods or services 

are not sold through markets, making it more difficult to obtain information on the "price" of the 

service and the number of people enjoying the service (i.e., how many people benefit from the 

resource through a specific use). 

Ecological resources will also provide some services and ecological processes that indirectly 

benefit society. For example, a coastal wetland provides services as a wildlife habitat and fish 

nursery, as a means for flood control, and as a filtering system for run-off waters. These types of 

services, which are not bought and sold through markets, are referred to as indirect, non-market 

uses. Individuals may value these services even though they are not directly using the resource. 

Sometimes these types of services can be connected to other activities that humans value and, 

therefore, are valued through that relationship (see the discussion on identifying economic 

benefit endpoints in Chapter 2). 

Economists also recognize several different categories of non-use values. As the term implies, 

non-use values represent the value that an individual places on the ecological resource that does 

not depend on the individual's current use of the resource. Existence value, for example, refers 

to the value people place on knowing that a particular resource exists, even if they have no 

expectation of using the resource. Other examples of non-use values include bequest value, 

which refers to the value people place on a maintaining a resource for future generations, and 

altruism, or the value people place on maintaining resources that are important to their family 

and friends. 

As described in Chapter 5, the benefits of an action that improves a specific ecological resource 

can be estimated by estimating people’s willingness-to-pay (WTP) for improvements to the 

various types of goods and services provided by the resource. For example, in estimating the 

benefits of an action to improve the quality of a wetland area, one might consider that the 

wetland area serves as a primary breeding area for several species of birds. Therefore, one might 

estimate the change in the value of bird watching and recreational fowl hunting to the individuals 

using the area. To capture the total value or benefits of a change to a specific ecological 

resource, one also needs to consider the value of its role in supporting the ecosystem and the 

indirect benefits it provides to mankind. That is, one needs to also identify and evaluate the 

indirect, non-market uses and non-use values associated with an ecological resource. 

Page 99 
The economic benefit analysis should identify as many different goods and services (and values) 

affected by the policy or action. For example, if a policy is expected to improve ecological 

resources that support various bird populations, the economic benefit analysis might consider 

potential impacts on the following goods and services society derives from birds: 

C Food source (direct, market use); 

C Hunting, bird watching, and contributing to the aesthetic environment for hikers, 

campers, anglers, and other recreationists (direct, non-market use); 

C Component to an ecosystem that supports or provides other goods and services and 

contribute to maintaining biodiversity (indirect, non-market use); and 

C As an endangered species or to maintain the bird species for future generations (non-use 

value). 

The following four subsections elaborate on the types of goods and services that might be 

provided by an ecological resource and identify the economic techniques that might be 

appropriate for estimating the economic value of changes to these goods and services. 

Page 100 
6.2.1 Direct, Market Uses 

Direct, market uses refer to those goods and services provided by an ecological resource that are 

directly used by society and are bought and sold through the market system. Direct, market uses 

primarily refer to those goods produced by an ecological resource that are consumed by humans 

or serve as inputs in the production of other goods, such as food products, water, fuel sources, 

and building materials. Prices and quantities produced for these goods and services are directly 

observable. 

For example, one benefit of a policy to improve air quality might be measured through the value 

(i.e., change in welfare) of the increased productivity of commercial crops and timber 

production. Similarly, the benefit of an action to improve water quality might be measured 

through the value of the increased production of a commercial fishery (i.e., more fish caught and 

sold). 

It is important to remember, however, that the change in value of the direct, market uses (e.g., 

timber, crops, or fish) provided by an ecological resource (e.g., air, water) may represent only a 

portion of the total benefits of the change experienced by the ecological resource. 

Examples of Direct, Market Uses Provided by Ecological Resources 

‘ Food Source 

C Fish (specific species) -- commercial fishery 

C Crops (specific type: corn, beans, apples, etc.) -- commercial and home 

production 

C Animal (fowl, deer, etc.) -- commercial consumption 

‘ Building Materials 

C Timber (specific species) 

C Stone 

Page 101 
Page 102
‘ Fuel 

C Timber (specific species) 

C Coal 

C Oil 

‘ Drinking Water Supply 

C Ground water reservoir 

C Surface water reservoir 

‘ Medicine 

‘ Chemicals/Minerals 

Valuing Direct Market Uses 

There are a number of market-based approaches that may be useful in estimating the value of 

changes to a direct market use provided by an ecological resource. In most cases, a marketbased approach is used to estimate the demand and supply functions for the good or service. For 

some market goods, such as commodity crops and timber, detailed general and partial 

equilibrium models have been developed, which estimate demand and/or supply responses to 

changes in productivity, prices, and other variables. Impacts or changes to the ecological 

resource that affect the quantity or quality of the goods and services provided by the resource can 

be measured by estimating the change in the demand and supply functions resulting from the 

change and measuring the welfare change or change in willingness-to-pay. 

For relatively small changes that do not change the supply or demand for the good or service 

provided by the resource, the change in the value of the goods and services provided by the 

resource can be measured based on the increase (or decrease) in the quantity of the good or 

service provided and the market price of the good or service (see later section on estimating 

benefits using market price and supply/demand relationships for additional discussion of this 

issue). Other market-based valuation approaches, such as examining the cost of alternatives or 

the spending to provide similar goods or services, may also be useful when price or quantity 

information is not readily available. Although these second-best approaches can provide an 

estimate of the magnitude of the potential benefits, they do not directly reflect welfare changes. 

Specific techniques that can be used to value changes in market-based goods include: 

C Market-Price and Supply/Demand Relationships 

C Market-Based Valuation Approaches. 

Page 103 
6.2.2 Direct Non-Market Uses 

Direct, non-market uses of an ecological resource include those goods and services that are 

directly observed and used by humans, but are not sold or traded through an open, competitive 

market. Direct, non-market uses include both consumptive uses (e.g., recreational fishing and 

hunting) as well as non-consumptive uses (e.g., bird watching or boating). Direct, non-market 

uses are generally considered quasi-public/quasi-private goods because access or use of the 

resource can be controlled but is often not strictly regulated and the benefit or value to one 

individual does not affect the benefit or value to others up to a point (i.e., congestion reduces the 

benefit/value to all users). 

Examples of Direct, Non-Market Uses Provided by Ecological Resources 

‘ Fishing 

C Recreational Fishing (specific species, area) 

C Subsistence Fishing (specific species, area) 

‘ Beach Use (sunbathing, swimming, walking) 

‘ Recreational Hunting (specific species) - for sport and/or personal consumption 

‘ Bird Watching (general, specific species) 

‘ Tourism 

‘ Boating 

‘ Hiking/Camping 

‘ Animal Viewing, Photography, Feeding (general, specific species) 

‘ Sightseeing 

‘ Aesthetic Value 

Page 104 
Valuing Direct, Non-Market Uses 

These types of services are not bought and sold through observable markets and therefore, do not 

have market prices associated with their use. For most of these types of goods and services, 

however, the change in the quantity and/or quality of the service being provided is quantifiable 

(e.g., increased number of fish caught per fishing trip, increased number of beach or boating 

days, increased chance of viewing wildlife). Because these types of goods and services do not 

have market prices, non-market valuation techniques must be used to estimate the implicit prices 

for the goods and services provided by the resource. Some methods rely on the explicit 

transactions (e.g., entrance or licensing fees, spending to protect a resource) or observed choices 

that people make (e.g., travel decisions, home location) that are associated with the use of the 

goods and services provided by the ecological resource. These methods assume that people 

demonstrate, or reveal, the value they place on a good or service through the choices they make. 

Other methods rely on the responses of individuals using the resource to proposed choices or 

questions regarding the value they place on their use of the resource. In some cases, more 

sophisticated techniques and models, which combine information on engineering and biophysical 

processes with economic information, are used to estimate ecosystem changes and impacts to 

specific uses or services. 

Specific methods that may be useful in valuing changes to direct, non-market uses include: 

Revealed Preference Methods: 

C Hedonic Price Methodologies 

C Travel Cost Methodologies 

C Random Utility Models 

Stated Preference Methods: 

C Contingent Valuation 

C Contingent Activity and Combining Contingent Valuation with Other Approaches 

C Conjoint Analysis and Contingent Ranking. 

Page 105 
6.2.3 Indirect, Non-Market Uses 

Indirect, non-market uses of an ecological resource include those goods and services that provide 

an observable benefit to mankind but are not directly consumed or used by individuals. Indirect, 

non-market uses include many ecological processes that indirectly benefit mankind by 

supporting other ecological resources, maintaining viable ecosystems, and protecting the local 

environment. Indirect, non-market goods and services are usually public in nature because 

access or use of the ecological resource cannot generally be excluded and any number of 

individuals can benefit from the use of the ecological resource through these services without 

reducing the benefits accruing to anyone else. These goods and services are not sold or traded 

through an open, competitive market, although a community may pay for replacement or 

substitute goods (often through taxes) that provide the same public services as provided by the 

ecological resource. 

Examples of Indirect, Non-Market Uses Provided by Ecological Resources 

‘ Flood Control 

‘ Storm Water Treatment 

‘ Ground Water Recharge 

‘ Climate Control 

‘ Pollution Mitigation 

‘ Wave Buffering 

‘ Soil Generation 

‘ Nutrient Cycling 

‘ Habitat Value 

‘ Biodiversity 

Valuing Indirect, Non-Market Uses 

These types of services are not bought and sold through observable markets and therefore, do not 

have market prices associated with their use. Because these types of goods and services do not 

have market prices, non-market valuation techniques must be used to estimate the implicit prices 

for the goods and services provided by the resource. Some methods rely on the observed choices 

that people make that are related to the indirect, non-market goods and services provided by the 

resource. These methods assume that people demonstrate, or reveal, the value they place on the 

Page 106 
goods and services provided by ecological resources through the choices they make. In some 

cases, expenditures for replacement or substitute goods that provide the same public services as 

the ecological resource may be used to estimate the minimum value of the indirect, non-market 

services supported by the ecological resource. Other methods rely on the responses of 

individuals to proposed choices or questions regarding the value they place on the goods and 

services provided by the resource. 

Specific techniques that may be useful in estimating the value of changes to indirect, non-market 

uses include: 

Revealed Preference Methods: 

C Hedonic Price Methodologies 

C Replacement/Alternative Cost 

C Avoidance Expenditures 

Stated Preference Methods: 

C Contingent Valuation 

C Contingent Activity and Combining Contingent Valuation with Other Approaches 

C Conjoint Analysis and Contingent Ranking. 

Page 107 
6.2.4 Non-Market, Non-Use Values 

Non-market, non-use values of an ecological resource are the values that individuals hold for the 

resource unrelated to their current use of the goods and services provided by the resource. 

Individuals may value the existence of the ecological resource or the availability of the goods 

and services provided by the ecological resource although they do not directly consume or use 

the resource themselves. Non-market, non-use values may stem from the desire to ensure the 

availability of the resource for future generations, benevolence toward relatives and friends, 

sympathy for people and animals adversely affected by environmental degradation, or a sense of 

environmental responsibility. Additionally, the specific non-use values associated with a 

particular ecological resource may not be mutually exclusive: when asked directly, people are 

unlikely to be able to separately identify the non-use values they hold or distinguish between the 

value they place on direct or indirect uses and their non-use value(s). 

Examples of Non-Market Non-Use Values Provided by Ecological Resources 

‘ Scarcity Value 

‘ Option Value (although some consider this a use value) 

‘ Existence Value 

‘ Cultural/Historical Value 

‘ Intrinsic Value 

‘ Bequest Value 

‘ Altruistic Value 

‘ Philanthropic Value 

Valuing Non-Market, Non-Use Values 

These types of services are not bought and sold through observable markets and, therefore, do 

not have market prices associated with their use. Because these types of goods and services do 

not have market prices, non-market valuation techniques must be used to estimate the implicit 

prices for the goods and services provided by the resource. Furthermore, by definition, the nonuse value associated with an ecological resource cannot be estimated based on observed actions 

or choices made by individuals. Thus, to estimate non-use values economists must rely on 

people's responses to proposed choices or questions regarding the value they place on certain 

Page 108 
ecological resources (known as contingent valuation). Determining the total non-market, nonuse value associated with a change to an ecological resource is often difficult because the total 

value depends not only on the value each individual holds, but also on the appropriate number of 

such individuals to count in the valuation process. Additionally, as discussed in the later 

technique sections, the use of contingent valuation is very controversial and continues to be 

refined by economists, sociologists and psychologists. 

The following techniques are applicable for estimating non-market, non-use values: 

C Contingent Valuation 

C Contingent Ranking 

C Conjoint Analysis. 

Page 109 
6.3 APPROACHES TO MEASURING RESOURCE VALUES 

This section introduces the reader to the different types of approaches available to estimate the 

economic value (i.e., change in social welfare or willingness-to-pay) of a change in the quality 

and/or quantity of the goods and services provided by an ecological resource. Each valuation 

method has a different approach to eliciting the value that society places on such changes. Most, 

if not all, techniques require sophisticated econometric analysis to employ. 

This section organizes and explains the general types of valuation techniques and discusses, 

generally, what data might be required to implement each type of approach. A framework for 

understanding the similarities and differences between the techniques is presented, followed by a 

brief description of each category of techniques. 

More detailed descriptions of the individual techniques are provided in later sections. The 

information provided on the individual techniques is based on findings from the literature; the 

reader is encouraged to independently evaluate any technique for their own use. An additional 

reference document, EPA’s Guidelines for Preparing Economic Analyses (U.S. EPA, 2000), also 

reviews the various techniques available for benefits valuation. 

Valuation Techniques 

Valuation techniques can be grouped into four general categories according to the means by 

which preferences are revealed and the process by which these preferences are translated into 

monetary values (Mitchell and Carson, 1989; Freeman, 1993). To determine into which 

category a method falls, it is necessary to ask the following two questions: 

1. Does the technique use data or observations from people acting in real-world situations 

(i.e., revealed preferences) or from people responding to hypothetical situations (i.e., 

stated preferences)? 

2. Does the technique yield monetary values directly (i.e., direct estimation of willingnessto-pay) or must monetary values be inferred based on a model of individual behavior 

(i.e., indirect estimation of willingness-to-pay)? 

Exhibit 17 illustrates the matrix and the corresponding organization of the valuation techniques 

available for developing original valuation estimates (Mitchell and Carson, 1989; Freeman, 

1993). Benefits transfer analysis, which is not listed in the following table, relies on the results 

of previous analyses to develop a valuation estimate for a new policy case or study site. 

Following the table is a discussion of the four categories of approaches and benefits transfer 

analysis. 

Page 110 
Exhibit 17 

Categorization of Valuation Techniques 

Direct Estimation of WTP Indirect Estimation of WTP 

Revealed Preferences 

Approach 

Market Price/Quantity 

(Estimated Supply/Demand) 

Market Simulation Models 

Travel Cost Studies 

Random Utility Model 

Hedonic Studies 

User Fees 

Replacement Costs 

Avoidance Expenditures 

Referendum Voting 

Stated Preferences Contingent Valuation Studies Contingent Ranking 

Approach Contingent Activity 

Contingent Referendum 

Conjoint Analysis 

Note: Benefits Transfer Analysis relies on estimates developed using one or more of the 

techniques listed in this table. 

Direct, Revealed Preference Approaches 

Direct, revealed preference approaches require data on real-life choices made by individuals 

regarding their consumption or use of a particular good or service. These approaches assume 

that an individual who is free to choose the quantity of good or service they desire at a specific 

price will choose the quantity that maximizes their welfare (or benefits), given the constraints 

placed upon them by the market (e.g., limited individual income, availability of substitutes and 

other goods, limited availability of specific goods or services). Thus, these types of approaches 

can only be applied for goods and services bought and sold through markets. Competitive 

market prices and production cost information, for example, can be used to estimate supply and 

demand relationships, that can then be used to estimate the consumer and producer surplus 

associated with the goods or services provided by a resource. Alternatively, more complex 

market simulation models might be used to mimic market conditions in an effort to determine the 

value (or change in value) placed on a good or service. Estimating market relationships for a 

good or service requires, at a minimum, time series or cross-sectional data on the price of the 

good or service, the quantity sold and consumed, detailed cost and revenue information for 

representative producers, as well as data on the environmental change affecting the supply and/or 

demand for the marketed good or service. 

In some circumstances, market data may be useful in providing a lower bound estimate of the 

value of a good or service. User fees, or the amount paid to use the services provided by the 

resources at that site, indicate a lower bound for the value that individuals place on the use of a 

specific site. The replacement cost technique infers the value of goods and services from the 

cost of replacing the goods and services or of providing alternatives. 

Page 111 
Indirect, Revealed Preference Approaches 

Indirect, revealed preference approaches rely on the relationships between the value placed on a 

good or service not traded through markets that is affected by environmental quality and the 

other real-world choices that individuals make. These approaches typically require modeling of 

these relationships to infer values for the non-marketed good or service. Because of the need to 

model complex relationships in order to infer values for a specific good or service, these 

techniques tend to have fairly significant data needs, which may include: price and quantity 

information for consumption of related market goods and services; use or consumption 

information for the good or service one wants to value; characteristics of the goods or services as 

well as substitute goods and services; and characteristics of users. 

Travel cost studies, for example, have been used to estimate the value of a particular recreational 

activity, such as fishing, based on the time and expense required to partake in that activity. 

Similarly, in using the avoidance expenditures approach, the cost of a particular event (or 

benefits of preventing an event), such as flooding, is estimated based on current expenditures to 

prevent or reduce the negative impact of the event. Random utility models estimate recreational 

demand by focusing on an individual's choice among substitute sites for any given recreational 

trip. Hedonic property and wage models attempt to identify the value of environmental quality 

implicit in purchasers' willingness-to-pay for property and in the monetary value placed on 

working conditions, respectively. Referendum voting offers an individual a fixed quantity of a 

good or service at a fixed price. If the individual accepts the offer, it can be assumed that the 

person values the resource by at least that amount. Thus, referendum voting data (e.g., approval 

for new regulation or management scheme) can also be used to indicate the minimum value 

placed on protecting the resources affected by the outcome of the vote. 

Direct, Stated Preference Approaches 

Direct, stated preference approaches, or contingent valuation approaches, involve asking a 

sample group of people directly about the values they place on certain effects or changes. Some 

direct approaches used to determine an individual’s willingness-to-pay for a specific 

improvement include: 

• Asking each individual directly how much they would be willing to pay to ensure or 

prevent a change; 

• Asking each individual whether they would be willing to pay some specific amount of 

money to ensure or prevent a change, varying the amount of money across the sample; 

and 

• Conducting a bidding game with each individual to determine the maximum amount each 

would be willing to pay to ensure or prevent a change. 

By aggregating over the sample, an analyst can estimate a demand curve for the specific change, 

which can then be used to estimate total WTP for the change. Both the degree of environmental 

change and the cost of the change can be varied in a contingent valuation analysis. Contingent 

valuation analysis requires conducting a survey of a representative sample of individuals affected 

Page 112 
by the environmental change. Good survey design and implementation are critical to the success 

of a contingent valuation analysis. Unfortunately, these activities, as well as the analysis of the 

resulting data, are typically very time and resource intensive. 

Indirect, stated preference approaches 

Indirect, stated preference approaches are also contingent valuation studies, except that the 

individuals questioned are not asked directly about the value they place on a specific change. 

Instead, individuals are asked to make a decision about another situation that depends or 

otherwise relates to the value they would place on the specific change to be valued. The 

responses to these questions are then used to draw inferences about the value of changes to the 

non-market good or service of interest. For example, individuals may be asked: 

• To rank combinations of varying quantities or qualities of goods, including both market 

goods, which have prices associated with their use, and non-market goods, for which the 

analyst wants to estimate the value (Contingent Ranking); or 

• To estimate the change in their current level of activity or use of a specific good or 

service under alternative scenarios in which the availability and quality of the good or 

service is varied (Contingent Activity). 

Contingent ranking asks individuals to rank combinations of varying quantities and qualities of 

non-marketed environmental goods and services as well as other marketed goods. In a 

contingent activity study, individuals are asked hypothetical questions about their level of 

activity under alternative levels of availability and quality of an environmental good or service. 

In a contingent referendum study, respondents are asked whether they would vote yes or no for a 

policy or action that would impose a specific cost on them and provide or ensure a hypothetical 

quality or quantity of an environmental service. Values for the environmental goods or services 

are then inferred from the choices made by the individuals. Conjoint analysis uses data gathered 

from survey respondents concerning the relative importance of various features of a product to 

determine the willingness-to-pay for a particular feature. For any of these indirect, stated 

preference approaches, the data requirements and concerns will be the same as those associated 

with the direct, stated preference approaches. 

Benefits Transfer Approach 

Benefits transfer analysis can often be used to estimate the value of a particular change when the 

resources or time to conduct original research are not available. Benefits transfer is also a 

desirable approach in cases where good information already exists from previous studies of the 

good or service in question, particularly when studies exist for similar types of locations and 

resource users. This approach involves identifying other valuation studies of similar changes at 

similar sites and using, or transferring, the value from the previous study(ies) to the new site of 

concern. In some instances, additional data might be used to adjust the value estimate to better 

suit the new situation or to correct for errors introduced in the original study. More advanced 

benefits transfer analysis involves transferring a benefit function, demand function, or valuation 

model to a new study site. 

Page 113 
Data Sources 

In addition to selecting a valuation technique, it is also necessary to identify data sources that can 

be used in the valuation of public goods and services. Some of the data, such as the ecological 

components affected, will come from the ecological assessment. Other data will also need to be 

obtained from other sources. The type of data required depends upon which valuation technique 

is chosen. Data might include market data on the prices of various goods, data on the number of 

users (e.g., the number of fishermen using a specific fishery), the quantity used (e.g., acres of 

forests cut down in a given year’s lumber production), or some measure of the ecological 

resource itself (e.g., acres of wetlands). The individual valuation technique sections provide a 

detailed discussion of the types of data required to implement each technique. 

References and Further Reading 

Braden, J.B. and C.D. Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. 

North-Holland, Amsterdam: Elsevier Science Publishers. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Mitchell, R.C. and R.T. Carson. 1989. Using Surveys to Value Public Goods: The Contingent 

Valuation Method. Washington, D.C.: Resources for the Future. 

Page 114 
R ev e aled 

P referen ce 

6.3.1 Market Price and Supply/Demand Relationships A pproach 

The "market value" of a good or service that is conveyed through the market system is the price 

placed on the good or service. The price of a good represents the value of an additional unit of 

that good, assuming the good is sold through an undistorted, perfectly competitive market (i.e., a 

market with properly assigned property rights, full information, and no taxes or subsidies). 

Market prices can be used to value small changes in the quantity of a good or service being 

provided (i.e., small effects or changes that do not affect the supply of or demand for the product 

or service). For example, the value of increased commercial fish harvest in a specific bay could 

be estimated based on the market value of the additional fish caught (i.e., pounds of additional 

fish caught multiplied by market price per pound of fish), assuming that the increased harvest for 

the area under study will not affect the market price. 

The value (i.e., cost or benefit) of larger scale changes that are likely to affect the supply or 

demand for a good or service cannot be correctly valued using market prices. Using market 

price ignores the change in the extra value provided by the good or service to consumers (e.g., 

the amount consumers would be willing to pay above the market price, known as consumer 

surplus). For the same reason, the change in the total consumer expenditures for a good or 

service (market price times the quantity purchased) is generally not a good indicator of the 

benefits associated with a change in the use of that good or service. Where such a bias matters, 

other approaches are necessary for estimating the benefits or the change in willingness-to-pay 

resulting from a change in the goods or services provided by an ecological resource. 

Estimating Supply and Demand Relationships 

One approach is to estimate the supply and demand relationships for each service or product 

before and after the environmental change to estimate the benefits of a specific action. 

Depending on the good or service considered, the change to the ecological resource will cause a 

shift in the supply curve or the demand curve. The change in the willingness-to-pay, or benefits, 

associated with the action can then be estimated based on the change in the area above the 

supply curve and below the demand curve. The demand and supply curves, or functions, are 

estimated using past data on prices and quantities of the good sold, the cost of production inputs, 

and information on production relationships (i.e., the quantity of output produced with a given 

amount of inputs). 

Page 115 
Market Simulation Models 

Economists have developed market simulation models that combine economic, 

engineering, and biophysical information to estimate changes in market supply and/or 

demand relationships, and thus, the benefits, of an environmental change. Such models 

can be used to examine the relationship between changes in environmental quality, such as 

the amount of acid deposition, and “material damage,” including reductions in stocks of 

physical assets such as buildings, bridges, roads, and art, or changes in biological outputs, 

such as agriculture and vegetation. Environmental changes that affect the level of output 

or production will affect the price and quantity of the good on the market that can lead to 

further changes in output or production. Although simple estimates of changes in supply 

and demand relationships can be used to estimate the initial change in price and quantity, a 

more complex market simulation model is needed to estimate further changes that result 

from market interactions and feedback relationships. Market simulation models are 

regularly used to estimate the effects of changes in environmental quality on agricultural 

and timber production. Simulation models have also been used in material damage 

assessments to identify changes in production and consumption caused by environmental 

changes, identify the responses of input and output to these changes, and identify the 

adaptations affected factors can make to minimize losses or maximize gains from changes 

in opportunities and prices (Adams and Crocker, 1991). 

Valuing the benefits of a change to an ecological resource based only on a single or a few market 

goods or services provided by that resource is unlikely to capture the full benefits of the change 

because many other services provided by the resource that are not sold through markets may also 

be affected. In the case of an action that improves the quality of a forest, for example, the forest 

will provide improved habitat for other species of flora and fauna and better scenic views and 

recreational opportunities, in addition to the increased value of the forest as a supply of timber. 

Therefore, when using changes to market goods and services to estimate benefits, one should 

also consider the potential benefits associated with additional services provided by the resource 

that are not sold through markets. 

Advantages 

' For established markets, price, quantity, and input cost information should be readily 

available. 

' Actual consumer preferences are measured using observed data. 

Disadvantages 

' Market data may only be available for a limited number of goods and services provided 

by an ecological resource and may not reflect the value of all productive uses of a 

resource. 

Page 116 
' It may be difficult to correctly estimate demand and/or supply relationships if limited 

data on prices and quantities are available. 

' It may be difficult to separate the supply and demand effects and to isolate the effects of 

the environmental change. 

' Market-based analyses do not capture non-use value. 

Data Requirements 

This technique requires time series data on market prices for the resource, the quantity sold and 

consumed, and detailed cost and revenue information for representative producers, as well as 

environmental data for both before and after the change. 

References and Further Reading 

Adams, R.M. and T.D. Crocker. 1991. "Materials Damages," in Braden, John B. and Charles D. 

Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. North-Holland, 

Amsterdam: Elsevier Science Publishers. 

Braden, J.B. and C.D. Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. 

North-Holland, Amsterdam: Elsevier Science Publishers. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Hanley, N. and C.L. Spash. 1993. Cost Benefit Analysis and the Environment. Brookfield, 

Vermont: Edward Elgar Publishing Limited. 

Just, R.E., D.L. Hueth, and A. Schmitz. 1982. Applied Welfare Economics and Public Policy. 

Englewood Cliffs, New Jersey: Prentice-Hall. 

Loomis, J.B. 1993. Integrated Public Lands Management: Principles and Applications to 

National Forests, Parks, Wildlife Refuges, and BLM Lands, Chapter 6, Applying Economic 

Efficiency Analysis in Practice: Principles of Benefit-Cost Analysis. New York, New York: 

Columbia University Press. 

Page 117 
R e v e aled 

P referen ce 

6.3.2 Market-Based Valuation Approaches A pproa c h 

Although the goods and services provided by an ecological resource may not be bought and sold 

through the market, there may be other market transactions occurring that provide information 

regarding the value of the environmental good or service under study. When estimating the 

value of specific goods or services, for example, it may be useful to look at other market 

transactions, such as fees paid for use of similar services or spending on projects or activities 

designed to provide similar goods or services. When estimating the value of changes to an 

ecological resource (or the goods and services it provides) it may be useful to consider the 

estimated cost of alternative actions undertaken to produce similar results or, alternatively, the 

level of spending to prevent or reduce the negative impacts resulting from damage to an 

ecological resource. 

Although these measures cannot generally be expected to provide an exact measure of the 

benefits of a change to an ecological resource, they can be useful in developing preliminary or 

order-of-magnitude estimates. This section describes how the cost of alternatives or 

replacements, avoidance expenditures, simulated markets, referendums, and user fees might be 

useful in estimating the benefits of improvements to ecological resources. 

Alternative/Replacement Costs 

The cost of providing or replacing the goods or services that an ecological resource could 

provide can be used to estimate the value of those goods and services and, in some cases, the 

benefits of an action to protect or restore that ecological resource. This approach is based on the 

concept of revealed preference: by choosing to undertake an action to provide or replace certain 

goods and services, society demonstrates (or reveals) that it values the goods and services 

provided by the ecological resource (and correspondingly value the resource itself) by at least as 

much as the cost of the project. In other words, it is assumed that if society invests in a project 

to provide similar services to those provided by an ecological resource, then the value of the 

services provided can be assumed to be at least as great as the dollar amount spent on the project. 

Therefore, the cost of the project might also be used to approximate a lower bound for the value 

of the ecological resource that provides the same services. Specific examples include: 

• Using the cost of building a retaining wall to estimate the value of wave buffering 

services provided by a wetland or coastal marsh area; 

• Using the cost of fish breeding and stocking programs to estimate the value of fish 

nursery services provided by estuaries or upland streams; or 

• Using the cost of constructing and operating a storm water filtration plant to estimate the 

value of water filtration by wetland areas. 

In using this approach, however, it is important to keep in mind that because the goods or 

services replaced probably represent only a portion of the full range of services provided by the 

ecological resource, this approach is likely to underestimate the benefits of an action to protect 

or restore the ecological resource. In addition, this approach should only be applied if the 

Page 118 
project has been implemented or if society has demonstrated its willingness-to-pay for the 

project in some other way (e.g., approved spending for the project). Otherwise, there is no 

indication that the value of the good or service provided by the ecological resource to the 

affected community is greater than the estimated cost of the project. 

In a similar context, the cost or estimated value of alternative approaches to achieving an 

environmental goal (e.g., reduced pollution levels) can be used to estimate the value of changes 

(most often improvements) to an ecological resource. Under this approach, the estimated 

benefits of one program designed to protect or improve an ecological resource would be used to 

estimate the benefits of a different program that is also intended to protect the same resource. 

For example, the value of reducing NOx emissions, in terms of reduced nitrification of surface 

water bodies, might be estimated based on the estimated benefits of reducing the flow of 

nutrients from non-point source run-off to surface water bodies (see also Benefits Transfer). 

The concept and approach discussed above is different from the restoration/replacement cost 

approach used commonly in Natural Resource Damage Assessments (NRDA) (and incorporated 

in damage assessment models developed by the Department of the Interior (DOI) and the 

National Oceanic and Atmospheric Administration (NOAA)). The NRDA 

restoration/replacement cost approach uses the cost to restore, rehabilitate, or replace the 

damaged natural resource, in addition to the value of lost uses during the period when the 

resource is damaged, to determine how much the polluter should pay in compensation. The 

problem with using the cost of restoration or replacement as a valuation technique is that there is 

no direct link between the cost of the restoration activities and the value of the services provided 

by a natural (ecological) resource that would be lost without restoration. As a result, the 

estimated cost to restore or replace the natural resource will likely bear little relationship to the 

true social value or change in the value of the resource. 

Avoidance Expenditures/Averting Behavior 

Averting behavior and defensive or avoidance expenditure analyses are more commonly applied 

in efforts to estimate the benefits of actions that protect or improve human health. However, 

such approaches also may be applicable in estimating the benefits of actions that improve the 

state of an ecological resource. This approach is also based on the concept of revealed 

preference: by choosing to undertake the action, society demonstrates (or reveals) that it values 

the resource or the improvement of the resource at least as much as the cost of the action 

designed to protect or improve the resource. Some argue that this approach is inconsistent 

because few environmental actions and regulations are based solely on benefit-cost comparisons 

(particularly at the national level). As a result, the cost of a protective action may actually 

exceed the benefits to society. It is probably more likely, however, that the cost of those actions 

already taken to protect an ecological resource will underestimate the benefits of a new action to 

improve or protect the resource. 

Using this approach to estimate the benefits of an action that protects an ecological resource, one 

might look at the expenditures by society to prevent or reduce the negative impacts to the 

resource as a measure of the value or benefits of that action. For example, the cost of alternative 

controls to reduce effluent emissions to a water body could be used to estimate the value or 

benefits of reducing pollutant concentrations in the water body. 

Page 119 
Bartik (1988) shows formally how changes in defensive expenditures by households to alleviate 

the negative effects of pollution can be used to estimate the benefits of reducing pollutants. 

Exhibit 18 presents some of the possible measures for estimating the benefits of reducing 

pollutant levels using defensive expenditures: 

Exhibit 18 

Estimating the Benefits of Reducing Pollutants Using Defensive Expenditures 

Pollutant Defensive Expenditure Measures 

Air Pollution Clean or repaint exterior of house; install air 

purifiers or new air conditioners; visit the 

doctor more frequently; move away from 

pollution source 

Water Pollution New well; bottled water; water purifiers; 

move away 

Hazardous Waste Similar to both water and air pollution 

depending upon medium by which hazardous 

waste affects households 

Noise Pollution Storm windows; thicker walls; move away 

Radon in well water Filter or aerate water; bottled water; increase 

house air ventilation; move away 

Radon in Soil Underneath House Ventilate crawlspace of house; seal 

foundation of house; use thicker concrete in 

basement; increase house air ventilation; 

move away 

Source: Bartik (1988). 

Referenda 

Referenda provide an institutional basis for asking individuals' preferences for certain goods and 

services and may provide a basis for estimating the value of a particular change. A typical 

referendum might ask voters if they are willing to pay a specified amount to support a program 

that increases the supply of a public good. The decision to vote "yes" is based on the individual 

voter's assessment of whether the added benefit of the program exceeds the added cost of the 

payment. One of three conditions must exist to use an actual referendum to value a good or 

service (Mitchell and Carson, 1989): 

• The same people must vote for different levels of the public good at a fixed tax level or 

for a fixed level of the good at different tax levels. For example, a situation where a 

referendum fails and the supporters modify it for the next election; 

• Different jurisdictions vote on the same level of a good; or 

• Different jurisdictions vote on different levels of a good. 

Page 120 
User Fees 

User fee information, such as entrance fees or other fee receipts, can be used to infer the value 

individuals place on the use of a specific site, such as a national park. User fees indicate a lower 

bound for the value that individuals place on the use of a specific site. At a minimum, it can be 

assumed that each visitor values their visit (or use) of the ecological resource by at least as much 

as they paid as an entrance fee or other charge to use the services provided by the ecological 

resource. User fee data alone, however, is likely to significantly underestimate WTP, because it 

misses values such as existence and option value and does not capture other “travel costs.” 

If one assumes that visitors react to increases in entrance fees in the same manner as to increases 

in travel costs, entrance parameters can be used to trace a demand curve for the site, much in the 

same way as under a travel cost study where the area under the demand curve is the measure of 

the value of the ecological resource. In addition, it is possible to use entrance fees or other 

charges assessed on users as a component of a broader travel cost or random utility model study. 

Simulated Markets 

Simulated market studies estimate what a person would pay for a good that is not sold on the 

market by creating market conditions for that good. Under market conditions, the price that a 

person will pay for a good or service is the value that the person places on that good or service. 

Therefore, by mimicking market conditions, one should be able to estimate the value that a 

person places on public goods and services. 

This technique can have advantages over other valuation methods, such as contingent valuation 

and travel cost. Like simulated market studies, these techniques attempt to attach value to public 

goods; however, they do not simulate market conditions, and therefore certain biases exist that 

affect their ability to estimate value. 

There may also be biases associated with simulated market studies, however, due to the 

potentially limited scope and artificial nature of the study. Additionally, conducting a simulated 

market study could be potentially costly. Simulated market studies may be most useful in 

limited contexts for interpreting the results and biases of contingent valuation, travel cost, and 

other valuation techniques (Bishop and Heberlein, 1979). 

Page 121 
Example Simulated Market Study (Bishop and Heberlein, 1979) 

This study used simulated markets, contingent valuation, and travel cost to estimate the value of 

goose hunting permits. Goose hunting permits were readily available -- hunters wrote in and 

requested permits for a specific season. Each permit allowed the hunter to take one goose and no 

fees were charged for the permits. Three samples of hunters were drawn from the total number 

of hunters who were issued permits. For the simulated market approach, the first sample of 

hunters received cash offers for their permits by mail; if the hunter accepted the offer, they were 

to send the permit back, otherwise, the check. The cash offers ranged between $1 and $200. A 

second sample of hunters received contingent valuation questionnaires in the mail designed to 

measure the value of the permits. The third sample received travel cost questionnaires designed 

to estimate a travel cost demand curve. 

Responses to cash offers yielded a total consumer surplus for the permits of $800,000 total, or 

$63 per permit. The contingent valuation survey estimated that the average willingness-to-sell 

was $101 per permit, while the average willingness-to-pay was $21 per permit. The travel cost 

study estimated costs per permit of $11, $28, or $45 based on the assumptions regarding the 

value of time (time value equals zero, 1/4 median income rate, and 1⁄2 median income rate, 

respectively). 

In theory, the simulated market study approximates the true value of the permit more closely 

than a contingent valuation study would because real money was used, and people were asked to 

make a choice similar to the market choices that are made each day. 

References and Further Reading 

Adams, R.M. and T.D. Crocker. 1991. "Materials Damages," in Braden, John B. and Charles D. 

Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. North-Holland, 

Amsterdam: Elsevier Science Publishers. 

Bartik, T.J. 1988. "Evaluating the Benefits of Non-marginal Reductions in Pollution Using 

Information on Defensive Expenditures." Journal of Environmental Economics and 

Management, 15: 111-127. 

Bishop, R.C. and T.A. Heberlein. 1979. "Measuring Values of Extramarket Goods: Are 

Indirect Measures Biased?" American Journal of Agricultural Economics, December: 926-929. 

Braden, J.B. and C.D. Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. 

North-Holland, Amsterdam: Elsevier Science Publishers. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Page 122 
Hanley, N. and C.L. Spash. 1993. Cost Benefit Analysis and the Environment. Brookfield, 

Vermont: Edward Elgar Publishing Limited. 

Just, R.E., D.L. Hueth, and A. Schmitz. 1982. Applied Welfare Economics and Public Policy. 

Englewood Cliffs, New Jersey: Prentice-Hall. 

Loomis, J.B. 1993. Integrated Public Lands Management: Principles and Applications to 

National Forests, Parks, Wildlife Refuges, and BLM Lands, Chapter 6, Applying Economic 

Efficiency Analysis in Practice: Principles of Benefit-Cost Analysis. New York, New York: 

Columbia University Press. 

Page 123 
Revealed 

Preference 

6.3.3 Travel Cost Methodologies Approach 

The travel cost method was developed as a technique to value public recreation sites. This 

technique incorporates the assumption that individuals visiting a recreational site pay an implicit 

price for the site’s services that includes the cost of travel to the site and the time spent visiting 

the site. Travel cost models pay special attention to the value of time. 

To illustrate the concept behind travel cost models, consider, for example, that on a particular 

day a person chooses to go to work or to a park (or engage in some other activity). The person 

must first decide whether or not to go to work and, if the person decides to go to the park, he or 

she must decide how much time to spend there. The cost of the visit to the park includes the cost 

of getting to the park, any entry fee that is paid, plus the foregone earnings, or opportunity cost, 

one could have earned by going to work. If these costs and the number of trips made in one 

season were assembled for a large population, the unit willingness-to-pay for a certain number of 

visits could be estimated (Pearce and Turner, 1990). 

In calculating the average willingness-to-pay for a trip using the travel cost method, it is 

important to note factors that require careful attention. In determining the number of trips taken 

by individuals, it is necessary to recognize that some visits may be multi-purpose trips and some 

trips may be taken by holiday-makers while others may be taken by residents. Furthermore, it 

may be difficult to accurately calculate distance costs and the value of time associated with 

visiting the site. These factors may require special attention in order to accurately estimate the 

value of the resources at the site (Hanley and Spash, 1993). 

To determine a demand curve for recreation at a specific site, it is necessary to understand that 

trip costs are like prices. Theory dictates that if prices are lower, people will consume a higher 

quantity of the good, or, in this case, if trip costs are lower, people will take more trips to the 

site. By plotting trip cost against the number of trips to the recreation site from different areas, a 

demand-curve for recreation days can be traced (Loomis, 1993). 

Typically, travel cost models are used to estimate the demand curve for an individual, although 

aggregate or market demand for a site might also be modeled. The consumer surplus for an 

individual visitor is the area under the estimated demand curve but above the trip cost. Because 

people come from different distances to use the site, consumer surplus is different for each user. 

People living close to the site "buy" more trips and pay less per trip. Hence, these people receive 

a much larger consumer surplus than people living farther from the site who "buy" fewer trips 

and pay more per trip. In other words, people living close to the site are willing to pay more than 

those living further away to have access to the site or to prevent deterioration of its 

environmental quality. Total consumer surplus for all individuals is found by adding up all of 

the trips from all locations and adding up each individual's consumer surplus. The average 

consumer surplus per trip can be used as an estimate of the average willingness-to-pay for a trip 

(Loomis, 1993). 

Page 124 
Shifts in the demand curve due to an improvement in the quality of the site can be used to 

estimate the change in the value of the site, or the benefit resulting from the improvement.2 

Similarly, because environmental quality is expected to influence demand for a site, changes in 

visitation rates for sites with different levels of environmental quality, holding travel costs 

constant, can be used to estimate the benefit of changes in environmental quality (Pearce and 

Turner, 1990). However, the random utility model may be a more appropriate technique when 

examining the choice between multiple sites. 

Advantages 

' The travel cost method can provide benefit measures for changes in environmental 

quality from the observed behavior of visitors to recreation sites. 

' The method can be adapted to many environmental quality issues where changes in 

quality affect the desirability of a recreation site. 

' The method can be implemented using mail, phone, interview surveys, or site registration 

data. In some cases, data are available from state and federal resource management 

agencies. 

Disadvantages 

' Travel cost studies may over- or under-estimate the value of a good or service if they use 

an inappropriate estimate for the market price of the time that people spend traveling to a 

recreation site. Economists continue to disagree about whether the value of travel time 

should be based on the person's wage rate, some fraction of their wage rate, or valued at 

zero. 

' The method can provide benefits information only on changes in environmental quality 

that have a direct effect on the site preferences of recreationists. Quality characteristics 

that users are indifferent to or unaware of cannot be evaluated. 

' Exclusion of alternative recreation sites and their characteristics (environmental quality 

and other site features) from the travel cost model may bias the benefit estimates. 

' Excludes non-use values. 

' Environmental quality and other site characteristics may be difficult to describe in 

quantitative terms. 

Data Requirements 

2 Because the travel cost method does not provide for estimation of the theoretically correct 

measure of WTP for a site or for changes in the environmental quality at a site, such estimates should be 

used cautiously. Furthermore, because of this potential limitation, one might consider the appropriateness 

of utilizing a method of exact welfare measurement, where the functional form for the travel cost demand 

curve is derived from an explicit specification of the individual’s utility function (Freeman, 1993). 

Page 125 
Travel cost models typically have the following data needs: (1) the county of residence or zip 

code for users of the recreation site, population size, and summary measures for features of the 

population in each origin zone; (2) round-trip mileage to the site and to substitute sites; (3) mode 

of transportation; (4) vehicle operating costs per mile and implicit time cost of travel; and (5) 

data on on-site characteristics, such as size, number, location, and type of facilities. Typically, 

this information is collected through surveys using phone, on-site, or mail surveys, or by 

acquiring site registration data. 

Example Travel Cost Study (Bockstael et al., 1989) 

A travel cost model was used to estimate the value of improved water quality to Maryland beach 

users on the western shore of the Chesapeake Bay. Water quality was measured as the product 

of the concentrations of nitrogen and phosphorous in the water at the monitoring site nearest to 

the beach in question. Data for the model was obtained from a survey of 484 people at 11 public 

beaches in the study area. 

The model was used to calculate the willingness to pay for a 20 percent improvement in water 

quality -- that is, a 20 percent reduction on total nitrogen and phosphorus. The average annual 

aggregate benefits to beach users of water quality improvement were estimated to be $35 million 

(1984 dollars). The long-run benefits to beach users of water quality improvements may be 

higher than the estimates reported, however, for several reasons. First, as people learn that the 

Bay has become cleaner, they will adjust their preferences toward beach recreation. People who 

do not currently use the Bay beaches will be especially likely to make this change. Additionally, 

the population and income of the area have grown and are likely to continue growing, increasing 

the demand for and value of the water quality improvements. Finally, the estimates given ignore 

households outside the Baltimore-Washington Statistical Metropolitan Sampling Area. 

References and Further Reading 

Bishop, R.C. and T.A. Heberlein. 1979. "Measuring Values of Extramarket Goods: Are 

Indirect Measures Biased?" American Journal of Agricultural Economics, December: 926-929. 

Bockstael, N.E., K.E. McConnell, and I.E. Strand. 1989. "Measuring the Benefits of 

Improvements in Water Quality: The Chesapeake Bay." Marine Resource Economics 6(1): 1­

18. 

Bockstael, N.E., K.E. McConnell, and I.E. Strand. 1991. “Recreation.” in Braden, J.B. and C.D. 

Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. North-Holland, 

Amsterdam: Elsevier Science Publishers. 

Fletcher, J., W. Adamowicz, and T. Graham-Tomasi. 1990. “The Travel Cost Model of 

Recreation Demand.” Leisure Sciences 12: 119-147. 

Page 126 
Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Hanley, N. and C. Spash. 1993. Cost Benefit Analysis and the Environment. Brookfield, 

Vermont: Edward Elgar Publishing. 

Loomis, J.B. 1993. Integrated Public Lands Management: Principles and Applications to 

National Forests, Parks, Wildlife Refuges, and BLM Lands, Chapter 6, Applying Economic 

Efficiency Analysis in Practice: Principles of Benefit-Cost Analysis. New York: Columbia 

University Press. 

McConnell K. and I. Strand. 1981. “Measuring the Cost of Time in Recreation Demand 

Analysis.” American Journal of Agricultural Economics: 153-156. 

Pearce, D.W. and R.K. Turner. 1990. Economics of Natural Resources and the Environment. 

Maryland: The Johns Hopkins University Press. 

Willis, K. and G. Garrod. 1991. “An Individual Travel Cost Method of Evaluating Forest 

Recreation.” Journal of Agricultural Economics 42(1): 33-42. 

Page 127 
Revealed 

Preference 

Approach 6.3.4 Random Utility Model 

The random utility model is a popular method to estimate consumers’ recreational demand. The 

random utility model is also known as a “discrete choice model” because it is used to study 

people’s choices between one or more alternatives. The term “random” refers to the fact that the 

model cannot directly observe people’s decision processes. The economist observes the final 

decision but must assume the decision process is logical, with people choosing the alternative 

providing the greatest possible level of satisfaction. The lack of direct observation is what 

makes the process “random” to an economist. 

With respect to valuing changes to ecological resources, the use of random utility models 

focuses on the choices individuals make among substitute sites for any given recreational trip 

rather than the number of trips a recreationist takes to a given site in a season, as with the travel 

cost model. The random utility model is especially suitable when the selection of alternatives or 

substitutes differ in terms of their quality or other characteristics. The random utility model is 

particularly appropriate when there are many substitutes available and when the change being 

valued is a change in a specific quality characteristic of one or more sites, such as catch rates or 

water quality. The random utility model can also be used to value the benefits of introducing a 

new site (U.S. EPA, 1995). 

The characteristics of the alternative sites that are used in the estimation of the model are 

instrumental in explaining how people allocate their trips across sites. Sometimes information 

on the characteristics of the individuals making the choices are also used in estimating a random 

utility model. 

Advantages 

' The random utility model can provide benefit measures for changes in environmental 

quality from the observed behavior of visitors to recreation sites. 

' The method can be adapted to many environmental quality issues where changes in 

quality affect the desirability of a recreation site. 

' This technique is preferred over the travel cost model for handling the issues of substitute 

sites and environmental quality considerations. 

Disadvantages 

' An inappropriate estimate for the value of time that people spend traveling to a site can 

adversely affect the estimated value of the good or service. 

' The method can provide benefits information only on changes in environmental quality 

that have a direct effect on the site preferences of recreationists. Quality characteristics 

that users are indifferent to or unaware of cannot be evaluated. 

' Model specification, as with all techniques and estimation procedures, can have a 

significant impact on benefit estimates. 

Page 128 
' This technique requires a significant amount of data. 

' Benefit estimates may be biased if: (1) known substitute sites are not included in the 

model or (2) additional substitute sites are included in the model that are unknown to the 

individuals surveyed. 

' Excludes non-use values. 

Data Requirements 

The random utility model has data needs similar to those of the travel cost model, including the 

cost of travel to the site or information to estimate the cost (i.e., distance traveled, any fees paid, 

plus the value of the individual's time) and characteristics of the chosen site and alternative sites. 

In addition, the researcher needs to know what alternative sites are considered by recreationists 

and may want to collect information on the characteristics of the individuals (e.g., education, 

income, other socio-demographic information). Additionally, accurate measurement of the 

characteristics of the alternative sites is necessary. 

Example Random Utility Study (Englin et al., 1991) 

This study uses both the random utility model and the travel cost method to estimate the 

damages to recreational trout fishing in the Upper Northeast due to acidic deposition. Data were 

collected on freshwater recreational trips made during the summer of 1989 by 5,724 randomly 

selected individuals in four Northeastern states: Maine, New Hampshire, New York, and 

Vermont. Changes in acidic deposition were expected to impact fish populations by changing 

acidic stress levels, thereby changing catch rates of various species. An angler's well-being 

should change when a change in the catch rate causes him or her to enjoy a site less (more) or 

results in a decision to change sites and travel farther (closer). The two models are based on the 

premise that the cost of travel to a site can be used to represent the price of a recreational fishing 

site. 

The random utility model provides estimates of changes in the value per choice occasion based 

on the relevant changes in the quality characteristics of the sites available to anglers. The model 

estimates that damages from acidic deposition are approximately $0.12 per trip. The travel cost 

model estimates the willingness to pay for a marginal increase in each attribute. With this 

technique, the willingness to pay for no damages from acidification was found to be $0.02 per 

trip. 

Page 129 
References and Further Reading 

Bockstael, N.E., K.E. McConnell, and I. Strand. 1989. “A Random Utility Model for 

Sportfishing: Some Preliminary Results for Florida.” Marine Resource Economics 6(1989): 

245-260. 

Englin, J.E., T.A. Cameron, R.E. Mendelsohn, G.A. Parsons, and S.A. Shankle. 1991. 

Valuation of Damages to Recreational Trout Fishing in the Upper Northeast due to Acidic 

Deposition. Richland, Washington: Pacific Northwest Laboratory. Prepared for National Acidic 

Precipitation Assessment Program. 

Hanemann, W.M. 1984 “Welfare Evaluations in Contingent Valuation Experiments with 

Discrete Responses.” American Journal of Agricultural Economics, August, 66: 332-341. 

Kaoru, Y., V. K. Smith, and J.L. Liu. 1995 “Using Random Utility Models to Estimate the 

Recreational Value of Estuarine Resources.” American Journal of Agricultural Economics, 

February, 77: 141-151. 

Smith, V.K. 1989 “Taking Stock of Progress with Travel Cost Recreation Demand Methods: 

Theory and Implementation.” Marine Resource Economics 6: 279-310. 

U.S. EPA, Oceans and Coastal Protection Division. 1995. Assessing the Economic Value of 

Estuary Resources and Resource Services in Comprehensive Conservation and Management 

Plan (CCMP) Planning and Implementation, A National Estuary Program Environmental 

Valuation Handbook. Washington, D.C. 

Page 130 
R e v ealed 

6.3.5 Hedonic Price and Hedonic Wage Methodologies P referen ce 

A p proach 

Hedonic methods typically use residential housing prices or labor wage rates, as well as other 

data, to measure the value of specific characteristics of a home, property, or job. These analyses 

identify the indirect linkage between environmental quality and the market price of a good or 

service, such as a residential property or employment opportunity, and use this linkage to 

estimate the implicit price, or benefit, of improvements in environmental quality. Under 

appropriate conditions, this implicit price can be interpreted as an individual's willingness-to-pay 

for environmental quality. In other situations, however, it is very difficult, if not impossible, to 

measure the welfare effects of a change to a specific characteristic, such as environmental 

quality. Nonetheless, the hedonic approach can still be useful for estimating a demand function 

for an environmental quality characteristic, such as the demand for proximity to a water body or 

distance from a hazardous waste facility. 

Hedonic Price 

The hedonic price, or property valuation, technique uses the assumption that the price of a house 

is a function of the characteristics of the home such as the quality of the surrounding 

neighborhood, the location of the home relative to business centers, and environmental 

characteristics including local air and water quality. The hedonic property model focuses on 

how changes in environmental quality affect property values by studying data from housing 

markets in different areas. Studying the relationships between changes in property values and 

differences in environmental quality can sometimes be used to determine an individual’s 

willingness-to-pay for improved environmental quality (Palmquist, 1991). The graph below 

illustrates the relationship between environmental quality and property values that might be 

uncovered by the hedonic property model. It shows that property values rise at a declining rate 

as the pollution level decreases or environmental quality improves. Other shapes of the hedonic 

function may be possible. 

Figure 1 

Graphic Illustration of a Hedonic Price Equation

 for an Environmental Quality Attribute (Pearce and Turner, 1990) 

Page 131 
When a change in environmental quality affects a large population, however, the hedonic 

property model alone may not be adequate to measure the change in welfare, or willingness-topay, and a more complicated analysis is required. In this case, some knowledge of the 

consumers’ preferences and a knowledge or a forecast of the change in the hedonic price 

equation (represented by PP’ in Figure 1) is necessary (Palmquist, 1991). A full discussion of 

this issue is beyond the scope of this document. 

In valuing changes in environmental quality, the hedonic approach attempts to do two things: 

• Identify how much of a property price differential is due to a particular environmental 

difference between properties; and 

• Infer how much people are willing to pay for an improvement in the environmental 

quality they experience (Pearce and Turner, 1990). 

For example, all other things being equal, one would expect prices of houses in neighborhoods 

with clean air to be higher than prices for houses in neighborhoods with polluted air. By 

comparing the market values of similar houses in areas with different levels of air quality, one 

may be able to determine that part of the difference in the price of housing in the two areas can 

serve as a measure of the value of clean air (Tietenberg, 1992). 

Example Hedonic Pricing Study (Palmquist et al., 1997) 

Palmquist, Roka, and Vukina used the hedonic pricing model to analyze the effects of hog 

operations on nearby houses. The authors examined the amount of hog manure located at 

varying distances from residential properties. Their purpose was to determine whether the 

presence of hogs influenced property values. 

Results from the hedonic model show that the presence of hog operations had a statistically 

significant negative effect on nearby property values. Changes in house values decreased as 

much as approximately $5,000 for a home located within 1⁄2 mile of a projected hog operation 

and as little as $1 for homes located 2 miles from the projected site in an area with higher 

concentrations of hog operations. The results indicated that the strongest negative impacts 

occurred closest to the hog operation and that effects on property value decreased as distance to 

the operations increased. Furthermore, in areas of high concentrations of hog operations, growth 

of hog operations experienced smaller negative effects than those areas with low concentrations 

of hog operations. 

Page 132 
Example Hedonic Pricing Study (Edwards and Anderson, 1984) 

Edwards and Anderson performed a hedonic price analysis that related the value of a house and 

its lot to characteristics of the house such as square footage, number of bathrooms, age, size of 

lot, and the following coastal zone characteristics: distance to a salt pond or ocean, frontage on a 

salt pond or ocean, and the presence of a view of the pond or ocean from the property. Their 

purpose was to determine the lost economic value to property owners associated with a zoning 

restriction in Narragansett Bay, Rhode Island. 

The results of the study suggest that the saltwater view and proximity to a salt pond are valued 

attributes of houses in the region. Using the estimated hedonic price equation, an approximate 

value of a water view was $5,790. It was further estimated that a land use policy restricting 

residential zoning in the salt pond region to protect groundwater supplies and water quality 

would result in lost opportunities for water view and water frontage locations valued at 

approximately $407,200. 

Hedonic Wage 

The hedonic wage technique is based on the presumption that, other things being equal, workers 

will prefer jobs with more pleasant working conditions. As the number of people seeking out the 

more pleasant jobs increases, the wages offered for such jobs will fall. Conversely, employers 

will have to raise the wage they offer for jobs with less pleasant working conditions to attract 

employees to these jobs. Therefore, at an equilibrium, the monetary value of better working 

conditions will be reflected in the difference in wages between two jobs with different working 

conditions (Freeman, 1993). 

Hedonic models are generally used to perform two types of valuations. The first, and more 

common, usage concerns the value of reducing the risk of death, injury or illness. In labor 

markets, workers that face higher levels of environmental or other job-related risk are 

compensated for that risk with higher than average wages. By estimating the dollar amount by 

which wages are increased to compensate workers for the greater risk, one can value the benefits 

that would be conferred by a reduction in the risk of death, injury, or illness (Tietenberg, 1992; 

Viscusi, 1993). Hedonic wage studies used to value the risk of illness or mortality may produce 

inaccurate results, however, if they do not account for the possible self-selection of less risk 

averse individuals into riskier jobs. 

Hedonic wage techniques can also be used, however, to value the environmental, social, and 

cultural amenities that vary across regions. This usage assumes that those cities and regions that 

are more desirable places to live and work in will attract workers from less desirable regions. As 

a result, employers in more desirable locations will pay lower wages, on average, than employers 

in less desirable locations for a worker with the same training and experience. Hedonic wage 

models try to measure the differences in wages between regions, or the “compensating wage 

differential,” to estimate the monetary value of differences in amenities (Freeman, 1993). 

Page 133 
Example Hedonic Wage Study (Bayless, 1982) 

Bayless used a hedonic wage analysis to relate the wages paid to academic professors and the air 

quality of the surrounding area. Bayless estimated a hedonic wage equation for salary of 

professors that incorporated pollution measures, as well as factors of productivity and locational 

characteristics. Bayless then used the hedonic wage equation to estimate a demand function for 

clean air, which was then used to estimate the willingness-to-pay for better air quality. 

This analysis found that the professors would be willing to pay approximately $100 to $400 per 

year in salary to move from areas of low air quality to high air quality. Willingness-to-pay 

values increased as the disparity in air quality between locations increased. 

Advantages 

' The hedonic techniques use market data on property sales prices and labor wages, these 

data are usually available through several sources and can be related to other secondary 

data sources to obtain descriptive variables for the analysis. 

' The technique is versatile and can be adapted to consider several possible interactions 

between market goods and environmental quality. 

' The hedonic method provides estimates of individuals' preferences for changes in 

environmental quality, which, under special conditions, can be interpreted as benefit 

measures. 

Disadvantages 

' The assumptions necessary to interpret the results of the hedonic technique as benefit 

measures are restrictive and, in many real world settings, implausible. Market 

equilibrium conditions require full knowledge of environmental effects that may be 

imperceptible to the physical senses. For example, if there are subtle long-term changes 

in water quality associated with some housing sites but people are unaware of the causal 

link of the effects to the housing site, their willingness-to-pay to avoid the effects will not 

be reflected in housing price differences. 

' Benefit estimates from a single product class will likely only capture a part of an 

individual's preferences for environmental quality. Property value models, for instance, 

are based on the consequences of individuals' choices of residence and therefore do not 

capture willingness-to-pay for improvements in environmental amenities at other points 

in the area, such as parks and recreational areas. 

' The estimating equations used for the hedonic technique may be statistically sensitive to 

model specification and estimation decisions. Appropriate tests for unbiasedness in 

housing and wage studies are still being developed. 

Page 134 
' Complete data on property or job characteristics may be difficult and expensive to gather, 

especially environment related characteristics. The omission of relevant characteristics 

and/or interactive environmental effects may reduce the validity of benefit estimates. 

' It may be difficult to isolate the specific amenity or environmental characteristic that is of 

interest. 

' Excludes non-use values. 

Data Requirements 

Data needs include sales or income, prices, wage data, characteristics of houses sold or jobs, and 

environmental amenity characteristics for each area of interest. These data can be collected from 

organizations such as multiple listing agencies, local tax assessors, and federal government 

agencies. Environmental quality data may be available from state, regional, or federal agencies 

and databases. Data collection, therefore, can often be time-consuming because of the effort 

required to gather data from a range of sources. The data sets can be gathered from markets that 

are separated either spatially or temporally or from a single market, although data from multiple 

markets tend to capture variation in price schedules, which may yield more accurate results. 

Additionally, while the data may be available, another problem may be the question of how 

individuals perceive their environment and whether individuals are aware of the quality of their 

environment. Most hedonic analysis use objective measures of environmental quality. 

However, some researchers have used subjective indicators, such as visibility, to determine 

environmental quality (Palmquist, 1991). 

References and Further Reading 

Bartik, T.J. 1988. “Measuring the Benefits of Amenity Improvements in Hedonic Price 

Models.” Land Economics 64(2): 172-183. 

Bayless, M. 1982. “Measuring the Benefits of Air Quality Improvements: A Hedonic Salary 

Approach.” Journal of Environmental Economics and Management 9(2): 81-99. 

Edwards, S.F. and G.D. Anderson. 1984. "Land Use Conflicts in Coastal Zone: An Approach 

for the Analysis of the Opportunity Costs of Protecting Coastal Resources." Journal of the 

Northeastern Agricultural Economics Council 13(1): 78-81. 

Freeman, A. M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Palmquist, R. 1991. “Hedonic Methods.” in Braden, John B. and Charles D. Kolstad, eds. 

Measuring the Demand for Environmental Quality. North-Holland, Amsterdam: Elsevier 

Science Publishers. 

Palmquist, R.B., F. Roka, and T. Vukina. 1997. “Hog Operations, Environmental Effects, and 

Residential Property Values.” Land Economics 73(1): 114-124. 

Page 135 
Pearce, D.W. and R.K. Turner. 1990. Economics of Natural Resources and the Environment. 

Maryland: The Johns Hopkins University Press. 

Tietenberg, T. 1992. Environmental and Natural Resource Economics. Harper Collins 

Publisher. 

Viscusi, W. K. 1993. “The Value of Risks to Life and Health.” Journal of Economic 

Literature. XXXI(4): 1912-1946. 

Page 136 
Stated 

Preference 

Approach 

6.3.6 Contingent Valuation 

Contingent valuation studies rely on surveys to ascertain respondent preferences for 

environmental goods and services by determining how much someone is willing to pay for 

changes in the quantity or quality of the good or service. These methods do not depend on 

market data; instead they establish a hypothetical market that gives survey respondents the 

opportunity to purchase the good or service. The dollar value that individual respondents are 

willing to pay for the good or service, when aggregated, can provide a means to value the good 

or service "sold" in the hypothetical market (Mitchell and Carson, 1989; Bateman and Willis, 

1998; Cummings et al., 1986). Because this method does not rely on market data, it can be 

applied to a variety of environmental quality issues for which market-based information is not 

available, including the elicitation of non-use values. 

Contingent valuation is a technique whereby people are asked what they are willing to pay for a 

benefit or what they are willing to receive by way of compensation to tolerate the loss of a good 

or service. The individual responses are aggregated to derive a demand curve for the good or 

service. 

A contingent valuation study is conducted either by written survey, interview, or some 

combination, and it generally consists of four parts: 

• Background information on the situation and possible changes to be made. 

• A detailed description of the good(s) or change to the good(s) being valued and the 

hypothetical method of payment. 

• Questions to elicit the respondents' willingness-to-pay for the good(s) or the change 

being valued. 

• Questions to collect socio-demographic (e.g., age, income); to validate the WTP response 

(e.g., what are their preferences relevant to the good(s) being valued, why did they give 

that dollar value); and to model their use of the good(s) (e.g., how frequently they visit 

the site). 

The aim of contingent valuation is to elicit valuations, or "bids," that are close to those that 

would be revealed if an actual market existed. The questioner, questionnaire, and respondent 

therefore, must represent as real a market as possible. For example, the respondent should be 

familiar with the good in question, such as improved scenic visibility, and with the hypothetical 

means of payment, such as a local tax or entry charge. 

Several individuals and groups have identified specific criteria for conducting reliable contingent 

valuation studies (Bjornstad and Kahn, 1996; Arrow et al., 1993; and Carson et al., 1996). 

Generally, these criteria include (at a minimum): 

• Interview a large sample of the affected population; 

• Achieve a high response rate; 

Page 137 
• Conduct in-person interviews when feasible; 

• Pre-test the questionnaire for interview effects and other potential biases; 

• Provide an accurate description of the event, program, or policy choice and the 

commodity to be valued; and 

• Remind interviewees of their budget constraints and the availability of comparable goods 

and services. 

While these guidelines are useful in assessing the reliability of a contingent valuation study, less 

restrictive and less costly approaches may be appropriate for informing policy decisions. In 

addition, some studies have found some of the above criteria (e.g., budget reminders) to have no 

effect on value estimates (Loomis, et al., 1994). 

Advantages 

' The contingent valuation method can be used to estimate the benefits of a variety of 

environmental effects for which market or secondary data are not available. 

' Comparisons of benefit measures from well-done contingent valuation studies with 

benefit estimates from other direct and indirect market techniques suggest that 

respondents can generally provide reasonable and consistent values for changes in 

environmental quality. 

' Contingent valuation methods are the only methods available for estimating non-use 

values (e.g., existence values). 

' The willingness-to-pay estimates from contingent valuation include both the use and 

nonuse value of the good or service. 

' Survey-based contingent valuation methods can capture respondents’ attitudinal and 

behavioral information that are not available in other non-survey based valuation 

techniques. 

' Useful for estimating non-use values. 

Disadvantages 

' The contingent valuation method is based on hypothetical situations in which it is 

difficult to verify whether expressed preferences are consistent with actual or planned 

behavior. Attempts to minimize the hypothetical nature of the process may only be partly 

successful. 

' Survey participants learn about their preferences for environmental quality during the 

valuation exercise. Survey design features may have a significant effect on this learning 

process and lead to responses that may not represent participants' true preferences. 

Page 138 
Conditional choice settings that are not at least partly familiar to the respondent may lead 

to uncertain responses. 

' Survey research is costly and time-consuming. National benefit estimates require 

properly designed sampling and enumeration procedures. Respondent refusals to 

consider environmental tradeoffs, as discussed in the choice exercises can raise questions 

regarding the validity of the benefit estimates. 

' Participants may answer strategically (high or low). Participants may provide 

unrealistically high answers if they believe that they will not have to pay for the good or 

service, but expect that their answer may influence the resulting supply of the good. This 

could lead to an overestimate of the actual willingness-to-pay. On the other hand, if 

participants believe that they might have to pay for the good or service based on the 

results of the survey, they might answer in such a way to keep the price low, and thereby 

cause surveyors to underestimate the value of the good or service. 

' Contingent valuation studies do not always find that WTP increases when the quality or 

quantity of a good or service increases. 

Example Contingent Valuation Study (Whittington et al., 1994) 

A contingent valuation survey was conducted of randomly selected households in the Greater 

Houston-Galveston Area to assess residents' willingness-to-pay for improvements in Galveston 

Bay's environmental quality and ecological health. In total, 234 interviews were completed in a 

mail/in-person follow-up survey, and 393 interviews in a mail-only survey. The analysis of 

responses showed that high-income respondents were more likely to vote for the management 

plan at a given price than low-income respondents; that users of the Bay were more likely to 

support the plan than passive users; and that people in general were less likely to vote for the 

management plan as the price of the plan presented as a monthly surcharge on their water bill 

increased. 

Based on the results of the mail-only contingent valuation survey, after adjusting results to 

account for differences between the socioeconomic profiles of respondents and the population of 

the study area, the authors estimated that the average household in the Greater HoustonGalveston Area is willing to pay approximately $7 per month, or about $80 per year, over five 

years for the management plan described in the questionnaire. 

Elicitation Methods 

There are several elicitation methods that are used in contingent valuation studies to determine 

an individual's willingness-to-pay. These methods represent different approaches for asking the 

respondent about their willingness-to-pay. 

Page 139 
The four methods discussed here include: 

• Direct, open-ended questioning 

• Payment card 

• Referendum/dichotomous choice 

• Iterative bidding games 

Each of these approaches is described below. 

Direct Open-Ended Questioning 

When using the direct open-ended questioning method, respondents are asked directly, 

“How much would you be willing to pay for the change in the good or service 

described?” 

Although the most obvious approach, it is also one of the most problematic methods. 

Advantages 

' Does not require pre-testing to determine an appropriate range for values as do the 

payment card and referendum voting methods. 

' Appears to provide conservative estimates of WTP. 

Disadvantages 

' Difficult for people to respond to an open-ended question because they are usually not 

accustomed to valuing environmental goods and services and typically do not face this 

type of question in a market situation. 

' May result in a high non-response rate and high number of extreme values (e.g., zeros 

and very large values). 

Payment Card 

The payment card method incorporates properties similar to the direct questioning approach but 

increases the response rates of willingness-to-pay questions. The payment card method asks the 

respondent to choose a willingness-to-pay amount from a card with a range of possible 

willingness-to-pay amounts usually starting from $0. The card sometimes indicates the average 

amount households of the same income range are willing to pay for other public goods (Mitchell 

and Carson, 1989). The payment card method, particularly with an average amount from other 

households, is no longer used in contingent valuation studies, but is described here for reference 

in reviewing older studies. 

Page 140 
Under this approach the respondent is asked: 

“What amount on this card or any amount in between is the most that you would be 

willing to pay for the level of good being proposed? 

Advantages 

' Provides more of a context for the respondent to provide a value. 

' Easier for respondent to select a value than to respond to an open-ended question. 

Disadvantages 

' Susceptible to biases associated with the ranges shown on card and the benchmark values 

provided by other households in the same income range. 

Referendum Voting/Dichotomous Choice 

Referendum voting/dichotomous choice is a technique where an individual is offered a fixed 

quantity of a good at a fixed price on a take-it-or-leave-it or yes-no basis. This is currently the 

favored approach for eliciting willingness-to-pay (WTP) estimates from survey respondents. 

Referendum voting as an elicitation method for contingent valuation differs from the use of 

actual referendum data described under Market-Based Valuation Approaches, in that a 

contingent valuation study referendum vote is a hypothetical scenario. While often referred to 

interchangeably, referendum style format and dichotomous choice can be distinct approaches. A 

survey could use a referendum scenario with more than two voting options (see example from 

contingent referendum section) and dichotomous choice could be used without a referendum 

scenario. Observing and analyzing the choices that individuals make through these techniques 

reveals the value of the good as it relates to the offered price (Freeman, 1993). For example, if 

someone accepts an offer to pay $10 a year in additional property taxes to preserve a wilderness 

area, it can be assumed that the person values the area by at least $10. If the resource was valued 

at less than $10, the person would not have accepted the $10 fee in a vote. However, the person 

may value the resource at more than $10 a year, and unless iterative voting is permitted, it would 

be impossible to determine the maximum that the voter is willing to pay. For this reason, 

referendum or dichotomous choice questions are often presented with one or two follow-up 

questions that present the respondent with a second choice scenario. This two-stage, or doublebounded, approach increases the statistical efficiency of the valuation estimate and reduces the 

necessary sample size. 

Page 141 
Advantages 

' Voting is a familiar social context, therefore respondents are likely to feel comfortable 

answering this type of question. 

' A vote provides a simple decision problem (either "yes" or "no"). 

' If the referendum questions are asked without an implied value judgment, there should be 

no starting point bias affecting the answers (Freeman, 1993). 

' Recent analysis has found the referendum question format to be strategic compatible (i.e., 

respondents are not expected to provide unrealistically high or low values for strategic 

purposes of supporting or suppressing the proposed (action). 

Disadvantages 

' Referendum voting requires more data and a larger sample size than direct questioning. 

' The outcome of referendum voting may be dependent on the distribution of offered bids, 

particularly the highest offered bid, because some respondents may be yea-saying or 

agreeing to pay any bid, no matter how high. 

' Outcomes of referendum voting may be dependent on the statistical methods used to 

analyze the responses. (See Haab and McConnell, 1997.) 

Example Contingent Valuation Study (Carson et al., 1996) 

A contingent valuation study using the referendum voting elicitation method was conducted by 

the National Opinion Research Center in 1993 of 1,182 residents in 12 U.S. cities to estimate the 

willingness-to-pay of individuals for a plan to provide two Coast Guard ships to escort oil 

tankers through the Prince William Sound to prevent future accidents and injuries due to oil 

spills. Willingness-to-pay was measured in terms of a one-time addition to Federal income 

taxes. During personal interviews, respondents were asked if they would be willing to pay a $10, 

$30, $60, or $120 one-time payment (each respondent was randomly assigned a dollar value). 

Based on the number of individuals willing to pay each dollar amount, the expected willingnessto-pay per individual was estimate to range from $50.61 to $52.81. 

Page 142 
Iterative Bidding Games 

Generally bidding games are conducted through personal interviews where the interviewer 

iteratively questions the respondent. Although this approach is generally no longer used because 

of bias issues, it is described here for reference in reviewing older studies. 

Questions are structured to lead to a "yes" or "no" response. For example, to estimate the value 

of environmental improvements, the interviewer might ask, 

"Would you continue to use this area if the cost to you was to increase by X 

dollars?" or 

"Would you be willing to pay an increase in your monthly electric bill of X dollars 

for Y reduction in air pollution?" 

The amount is varied with the same individual and the highest "yes" answer is recorded. 

Advantages 

' Able to get maximum willingness-to-pay from each individual surveyed. 

' May not require as large a sample as other approaches. 

Disadvantages 

' The outcomes of bidding games have been found to be highly dependent on the starting 

point, or first offered bid. 

' It can be difficult to develop a credible bidding game; the situation presented to survey 

respondents must be realistic and credible to the participants. Because of these 

difficulties, few recent contingent valuation surveys use bidding games to elicit values. 

Data Requirements 

The primary data for a contingent valuation analysis are acquired from a clearly defined and 

pretested survey of people who are representative of an affected population. A representative 

sample of the affected must be identified to allow extrapolation to the full affected population. 

Some additional research may also be required to determine the extent of the affected population 

or market for the good or service affected by the proposed action. 

The survey should generate data on respondents' willingness-to-pay for (or willingness-toaccept) a program or plan that will affect their well-being, as well as socio-demographic 

information and other data required to test for potential biases. A critical component of the data 

collection or survey implementation is the transfer of information to respondents about the 

resource, resource service, or action they are being asked to value. Photographs, verbal 

descriptions, video, and other multimedia techniques are commonly used to convey this 

information. In conducting a contingent valuation survey, the quality of the results depends in 

Page 143 
large part on the amount of information that is known beforehand about the way people think 

about the good or service in question. 

References and Further Reading 

Alberini, A. 1995. “Optimal Designs for Discrete Choice Contingent Valuation Surveys: 

Single-Bounded, Double-Bounded, and Bivariate Models.” Journal of Environmental 

Economics and Management, 28(3): 287-306. 

Arrow, K., R. Solow, P.R. Portney, E.E. Leamer, R. Radner, and H. Schuman. 1993. “Report of 

the NOAA Panel on Contingent Valuation.” Federal Register, January, Vol. 58(10): 4601-4614. 

Bateman, I.J. and K.G. Willis, Eds. 1998. Valuing Environmental Preferences: Theory and 

Practice of the Contingent Valuation Method in the U.S., E.U., and Developing Countries. 

Oxford University Press, Oxford. 

Bishop, R.C. and T.A. Heberlein. 1979. “Measuring Values of Extramarket Goods: Are 

Indirect Measures Biased?” American Journal of Agricultural Economics, December: 926-929. 

Bjornstad, D.J. and J.R. Kahn, eds. 1996. The Contingent Valuation of Environmental 

Resources: Methodological Issues and Research Needs. Brookfield, Vermont: Edgar Elgar 

Publishing Ltd. 

Carson, R.T. et al. 1996. “Was the NOAA Panel Correct About Contingent Valuation?” 

Washington, D.C.: Resources for the Future. 

Cooper, J.C. and J. Loomis. 1992. “Sensitivity of Willingness to Pay to Bid Design in 

Dichotomous Choice Contingent Valuation.” Land Economics, 68(2): 211-224. 

Cooper, J.C. 1993. “Optimal Bid Selection for Dichotomous Choice Contingent Valuation.” 

Journal of Environmental Economics and Management, 24(1): 25-40. 

Cummings, R.C., D. S. Brookshire, and W.D. Schulze, Eds. Valuing Environmental Goods: An 

Assessment of the Contingent Valuation Method. Rowan and Allanheld Publishers, Totowa, NJ. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Haab, T. C., and K. E. McConnell. 1997. “Referendum Models and Negative Willingness to 

Pay: Alternative Solutions.” Journal of Environmental Economics and Management, 32(2): 

251-270. 

Kannimen, B.J. 1993. “Design of Sequential Experiments for Contingent Valuation Studies.” 

Journal of Environmental Economics and Management, 25(1): s1-s11. 

Page 144 
Loomis, J., A. Gonzales-Caban, and R. Gregory. 1994. “Do Reminders of Substitutes and 

Budget Constraints Influence Contingent Valuation Estimates?” Land Economics. 70(4): 499-506. 

Mitchell, R.C. and R.T. Carson. 1984. An Experiment in Determining Willingness to Pay for 

National Water Quality Improvements. Washington D.C.: Resources for the Future. 

Mitchell, R.C. and R.T. Carson. 1986. The Use of Contingent Valuation Data for Benefit/Cost 

Analysis in Water Pollution Control. Washington D.C.: Resources for the Future. 

Mitchell, R.C. and R.T. Carson. 1989. Using Surveys to Value Public Goods: The Contingent 

Valuation Method. Washington, D.C.: Resources for the Future. 

Randall, A., B. Ives, and C. Eastman. 1974. "Bidding Games for Valuation of Aesthetic 

Environmental Improvements." Journal of Environmental Economics 1: 132-149. 

Whittington, D., et al. 1994. The Economic Value of Improving the Environmental Quality of 

Galveston Bay. Galveston National Estuary Program. Publication GBNEP-3B. 

Page 145 
Stated 

6.3.7 Combining Contingent Valuation with Other P referen ce 

Approaches: Contingent Activity A pproach 

In a contingent activity or contingent behavior study individuals are asked how they would 

change their behavior in response to a change in an environmental amenity. For example, one 

could use a contingent activity to estimate how a demand function for visits to a recreational site 

would shift with a change in one of the site's environmental attributes. Assuming that one has 

already estimated the demand for visits to a site under current conditions, the analyst asks 

visitors how their visitation behavior would change as a result of a change in an environmental 

attribute of the site (e.g., change in water quality). This information can then be used to estimate 

a shift in the demand curve for visits to the site (Freeman, 1993). 

In essence, a contingent activity approach combines the technique of contingent valuation with 

other valuation approaches used to model demand for a particular good or service to extend the 

application of these models to other scenarios. Recently, analysts have explored more advanced 

approaches for using travel cost data in combination with contingent valuation data to estimate a 

single joint model of individual’s preferences and demand for a particular good or service 

(Cameron, 1992). Future analysis is expected to also explore the use of travel cost information 

and contingent valuation responses to estimate a random utility or discrete-choice model. Jointly 

soliciting contingent valuation responses with other data, such as travel cost data or site-selection 

data, both (1) expands the ability of the model to account for both current users and non-users in 

characterizing demand and (2) lends credibility to the contingent valuation information. 

Advantages 

' Can expand the applicability of existing valuation analyses. 

' Potentially will allow for more complete characterization of demand by accounting for 

both current users of the resource and non-users. 

Disadvantages 

' The theoretical models and applied approaches for estimating demand using combined 

data are technically complex and not thoroughly developed. 

' It is not clear how to reconcile data from the different approaches if they do not 

correspond well. 

Page 146 
Example of Combining Contingent Valuation and Travel Cost Data (Cameron, 1992) 

The in-person survey, conducted by the Texas 

(1) 

(2) 

(3) 

restrictions on days of access. 

fishing days would result in a $35 loss in welfare, on average. 

In this study, Cameron combines contingent valuation responses and travel cost data on actual 

behavior collected through a single survey instrument to estimate a joint model of individual’s 

preferences and demand for fishing days. 

Department of Parks and Wildlife, asked 3,366 respondents: 

If they would have participated in salt-water fishing if their total annual cost was $X 

more, where the additional dollar amount was randomly chosen from $50 to $20,000; 

How much they will spend on their current fishing trip; and 

How many trips they took over the last year. 

The estimated model of demand for fishing days was then used to value greater and lesser 

Specifically, Cameron estimated that a 10 percent reduction in 

The complete loss of access was 

estimated to result in a $3,451 loss in welfare, on average. 

References and Further Reading 

Cameron, T.A. 1992. “Combining Contingent Valuation and Travel Cost Data for the Valuation 

of Nonmarket Goods.” Land Economics, August, 68(3): 302-17. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Mitchell, R.C. and R.T. Carson. 1989. Using Surveys to Value Public Goods: The Contingent 

Valuation Method. Washington, D.C.: Resources for the Future. 

Roe, B., K. Boyle and M. F. Teisl. 1996. “Using Conjoint Analysis to Derive Estimates of 

Compensating Variation.” Journal of Environmental Economics and Management, 31(2): 145­

159. 

Wittink, D.R. and P. Cattin. 1989. “Commercial Use of Conjoint Analysis: An Update.” 

Journal of Marketing, 53:91-96. 

Page 147 
Stated 

Preference 

Approach 

6.3.8 Conjoint Analysis and Contingent Ranking 

This section introduces the reader to conjoint analysis, a technique applied fairly recently to the 

valuation of environmental quality, and the more familiar approach of contingent ranking, which 

actually represents one type of conjoint analysis. 

Conjoint Analysis 

Conjoint analysis is a technique developed by marketing analysts used to value consumer 

preferences for specific features of goods or services. First, a composite good is separated into 

its constituent attributes. Then, individuals are surveyed regarding their relative preferences for 

alternative bundles of attributes, with multiple attributes varying simultaneously. The 

information gathered from survey respondents can then be used to calculate the marginal rates of 

substitution between the constituent attributes. By including price as one of the attributes, it is 

possible to rescale the utility index in dollars and derive estimates of willingness-to-pay for 

particular attribute bundles. 

Conjoint analyses generally fall into one of three types: ranking (same as contingent ranking 

approach discussed below), paired rating, and discrete choice. In a ranking study, respondents 

are often given several cards. Each card shows a unique product or program composed of 

specific attribute levels. Respondents are asked to put these cards in order -- from their most 

preferred to least preferred product or program. Alternatively, with the pairwise rating 

technique, respondents are shown two different products or programs simultaneously. 

Respondents are asked which product they prefer, and answer by supplying a rating within some 

range of number, for example, 1 to 9, where 1 indicates a strong preference for the first program, 

9 indicates a strong preference for the second program, and 5 indicates indifference between the 

two programs. Finally, the discrete choice technique provides respondents with several different 

products or programs simultaneously and simply asks them to identify the most-preferred 

alternative in the choice set. 

Conjoint analysis can be a useful technique in the valuation of improvements to ecological 

resources, given that several service flows are often affected simultaneously. For example, 

improved water quality in a lake will improve the quality of several services provided by the lake 

such as a cleaner drinking water supply, increased fishing and boating usage, and increased 

biodiversity. Conjoint analysis allows the valuation of these service flows both individually and 

as a whole. The technique also allows respondents to systematically evaluate trade-offs between 

multiple environmental attributes or between environmental and non-environmental attributes 

(Johnson et al., 1995). 

Page 148 
Example Conjoint Analysis Study (Mackenzie, 1992) 

This study develops a conjoint measure approach to evaluate unpriced attributes for 

recreational waterfowl hunting trips in Delaware. First, focus interviews were conducted 

with various hunters to identify major attributes of hunting trips that influence trip 

preferences. Four plausible levels were chosen for each of the following attributes: 

• Travel time (1, 2, 4, or 8 hours each way) 

• Trip cost per day ($25, $50, $100, or $200) 

• Type of hunting party (alone, with casual acquaintances, with close friends, or with 

family) 

• Site congestion (none, slight, moderate, heavy) 

• Hunting success (none, one duck, three ducks, three ducks and one goose) 

• Annual license fee (for state residents: $15, $20, $25, or $30; else: $45, $50, $60, or 

$80) 

A mailback survey questionnaire was designed to measure the relative preferences for these 

attributes by asking respondents to rank trip options providing alternative levels of each of 

the attributes. For example, respondents may have been asked to choose between (1) a trip 

with family to a slightly congested site two hours away, costing $100 per day, resulting in 

three ducks and requiring a $20 license, and (2) a trip with close friends to a heavily 

congested site one hour away, costing $25 per day, resulting in one duck and requiring a $15 

license. The survey was administered in 1989 to 3,351 hunters who purchased Delaware 

hunting licenses for the 1988-1989 hunting season. The survey generated 1,384 usable 

responses; of these, 696 respondents had hunted waterfowl at least one day during 

Delaware's 1988-1989 waterfowl season. 

A logistic model was then used to model these responses, and the marginal value of the 

various trip attributes could be calculated. The implied value of ducks bagged, for example, 

was found to be $81.35 per duck. The value of travel time was found to be $37.07 per hour. 

Advantages 

' Conjoint analysis allows the valuation of an action as a whole and the various attributes 

or effects of the action. 

' Respondents are allowed to systematically evaluate trade-offs among multiple attributes. 

' The trade-off process may encourage respondent introspection and facilitates consistency 

checks on response patterns (Johnson et al., 1995). 

' Respondents are generally more comfortable providing qualitative rankings or ratings of 

attribute bundles that include prices, rather than dollar valuations of the same bundles 

without prices. By de-emphasizing price as simply another attribute, the conjoint 

approach minimizes many of the biases that can arise in open-ended contingent valuation 

Page 149 
studies where respondents are presented with the unfamiliar and often unrealistic task of 

putting prices on non-market amenities (Mackenzie, 1992). 

' Because the technique has been so widely used in marketing literature, many of the 

statistical issues in the design and analysis of this type of study have been resolved. 

' Allows for questions regarding how one resource might be traded-off against another 

resource (rather than estimating WTP in terms of dollars). 

' Allows for the assessment of situations where some attributes of a resource improve 

while other attributes decline. 

Disadvantages 

' Respondents may find some trade-offs difficult to evaluate or unfamiliar to them. 

' A large number of trade-off questions may frustrate respondents. 

' Pairwise comparisons impose strict assumptions on preferences. 

' Although conjoint analysis has been used widely in the field of market research, its 

validity and reliability for valuing non-market commodities is largely untested (Johnson 

et al., 1995). 

' If the number of attributes or levels of the attributes is increased, the sample size and/or 

the number of comparisons each respondent makes must be increased. 

Contingent Ranking 

Contingent ranking asks respondents to hypothetically rank alternative choices or bundles of 

goods or services, where the alternatives vary in terms of their characteristics (e.g., representing 

different qualities or quantities of a good or service and different costs), in order of preference. 

These rankings can be analyzed to determine each respondent's preferences for the various 

attributes of the goods or services. If a monetary value can be assigned as one of the attributes, 

then it is possible to compute the respondent's willingness-to-pay for the environmental quality 

characteristic of the good or service on the basis of the ranking of the alternatives (Freeman, 

1993). 

One benefit of contingent ranking studies (compared to other contingent methods) is that 

respondents may be able to give more meaningful answers to questions about their behavior 

(e.g., they prefer one alternative over another) rather than to direct questions about the value of a 

good or service or the value of changes in environmental quality. The major challenge with 

contingent ranking is how to translate the answers into a dollar value. It may be necessary to 

imply a value from the relative ranking of other goods and services that do have a monetary 

value, which may lead to greater uncertainty in the actual value that is placed on the good or 

service of interest. 

Page 150 
For example, contingent ranking could be used to value a proposed change in the environmental 

quality of a recreational site. Respondents would be asked how they rank a set of sites that vary 

in two or more characteristics, where one characteristic is distance and another is level of 

environmental quality. Based on the ranking of the sites, the value of changes in environmental 

quality could be implied based on how distance (and therefore, the cost of travel) is traded off for 

other characteristics, including environmental quality (Mitchell and Carson, 1989). 

Advantages 

' Respondents may be more comfortable ranking alternative options rather than answering 

a willingness-to-pay question. 

Disadvantages 

' Contingent ranking requires more sophisticated statistical techniques to estimate WTP. 

' The respondents’ behavior underlying the results of a contingent ranking study is not 

well understood. 

' Contingent ranking tends to extract preferences in the form of attitudes instead of 

behavior intentions, and by only providing a limited number of options, it may force 

respondents to make choices that they would not voluntarily make (Mitchell and Carson, 

1989). 

References and Further Reading 

Desvousges, W., V.K. Smith, and M.P. McGivney. 1983. A Comparison of Alternative 

Approaches for Estimating Recreation and Related Benefits of Water Quality Improvement. 

Prepared for the U.S. EPA. EPA/230/05-83/001. Washington D.C.: U.S. EPA. 

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Washington, D.C.: Resources for the Future. 

Johnson, F. R., W.H. Desvousges, L.L. Wood, and E.E. Fries. 1995. Conjoint Analysis of 

Individual and Aggregate Environmental Preferences, Technical Paper No. T-9502. Triangle 

Economic Research. 

Mackenzie, J. 1992. “Evaluating Recreation Trip Attributes and Travel Time via Conjoint 

Analysis.” Journal of Leisure Research 24(2): 171-184. National Recreation and Park 

Association. 

Mitchell, R.C. and R.T. Carson. 1989. Using Surveys to Value Public Goods: The Contingent 

Valuation Method. Washington, D.C.: Resources for the Future. 

U.S. EPA, RTI. 1983. A Comparison of Alternative Approaches for Estimating Recreation and 

Related Benefits of Water Quality Improvements, EPA Document 230-05-83-001 Under 

Cooperative Agreement #68-01-5838. Washington D.C.: U.S. EPA. 

Page 151 
Page 152
B enefits 

Transfer 

A pproach 6.3.9 Benefits Transfer 

Benefits transfer is often used in benefit-cost analysis when limited resources or time constraints 

make it difficult to conduct an original valuation study. Benefits transfer involves obtaining an 

existing estimate of an economic use value (e.g., unit willingness-to-pay per individual) or 

demand function from a previous study to estimate the value associated with a similar use being 

provided by a similar ecological resource under another policy case or at a new study site. The 

benefit estimate from the original valuation study is scaled by the level of change under the new 

policy case or level of use at the new study site (e.g., number of users) to estimate the benefits of 

a similar change in the services provided by the ecological resource under study. 

Benefits transfer is most reliable when (U.S. EPA, 1995): 

• The original policy case or site and the new policy case or study site are very similar; 

• The environmental change is very similar for the original and new analyses; and 

• The original valuation study was carefully conducted and used sound valuation 

techniques. 

The reliability of the benefits estimate developed using the benefit transfer technique depends 

primarily on the similarity between the original and the new policy case or study site. With 

respect to benefits transfer between sites, large differences in quality, location, visitor 

characteristics, availability of substitutes, or object of valuation between the original and the new 

site have been found to impact the reliability of the benefit estimates derived through benefit 

transfer (Kirchhoff et al., 1997). 

There are three commonly used benefit transfer techniques: 

• Mean unit value transfer; 

• Adjusted unit value transfer; and 

• Benefit/demand function or model transfer. 

When possible, the transfer of demand functions or models is generally preferred to the use of 

unit value estimates. Both Loomis (1992) and Kirchhoff et al. (1997) conducted empirical 

analyses that found the transfer of a benefit or demand function was more reliable (e.g., smaller 

percentage errors) than a unit value transfer approach. 

Mean Unit Value Transfer 

Average unit values are generally used in benefits transfer analysis when either the demand 

function or model used for the original study is unavailable or the input data for a demand 

function or model is not available for the new policy case or study site. Average unit values are 

often used for regulatory analysis because the broad, typically regional or national, scope of the 

analyses makes it impossible, and often inappropriate, to reestimate a demand function or model 

Page 153 
developed for a specific location. The mean unit value technique assumes that the use value of a 

resource change under the original policy case or at the original site can be applied directly to the 

new policy case or site without adjustment. In this case, the unit value estimates generally apply 

to a specific use of the resource (e.g., recreational fishing, duck hunting, fresh water swimming) 

and represent an average or median value developed from a wide range of studies. 

Adjusted Unit Value Transfer 

The unit value estimate may be adjusted before it is applied to the new study situation to correct 

for any bias or inaccuracies associated with the original valuation study or to adjust for 

differences in the attributes of the policy case or study site that would affect the value estimate. 

Under the adjusted unit value technique, adjustments are generally made to account for three 

types of differences between the original and the new policy case or study site (U.S. EPA, 1995): 

• Differences in attributes of the policy case or site, level of use, or in the socioeconomic 

characteristics of users affected by the change; 

• Differences in the environmental policy, change, or resulting effects; and 

• Differences in the availability of substitute goods and services. 

Additional adjustments may also be made to the nominal value from the original study(ies) to 

update the value estimate to current year dollars. If the benefits transfer application is using 

multiple primary valuation studies from different study years, the estimates will need to be 

converted to the same year dollars to allow comparisons to be made. 

Benefit/Demand Function or Model Transfer 

A final option under benefits transfer is to transfer the entire demand function or valuation 

method estimated by another valuation study to the new policy case or study site. In most 

circumstances, as with transferring a unit value estimate, the demand function may need to be 

adjusted to better suit the characteristics of the new policy case or study site. The transferred 

demand function can then be used to estimate the willingness-to-pay or benefits associated with 

improving the service provided by the ecological resource. When the demand function is 

transferred, the benefit estimate captures both changes in the level of use and unit value benefit 

estimate for the new study site (Loomis, 1992). Recent research suggests that when conducting 

a benefits transfer analysis for a new study site, benefit or demand functions that account for a 

larger number of site characteristics may provide for more accurate benefit transfer analysis 

(Kirchhoff et al., 1997). Unfortunately, the use of more detailed benefit or demand functions 

increases the need to collect site-specific data for both the original study site and the new study 

site (or policy-specific data in case of a benefits transfer analysis for a new policy case), which 

increases the time and resource costs of benefits transfer analysis. 

Models for valuing ecological resources and damages to ecological resources can also be 

transferred in their entirety. Any valuation model being considered should be evaluated to 

determine its applicability to the new study situation, much in the same way as a unit value 

Page 154 
estimate or demand function must be reviewed for appropriateness before it is used to estimate 

the value of a change in a service under a different policy case or at a different site. 

Advantages 

' Economic benefits can be estimated more quickly than if undertaking an original 

valuation study. 

' Benefits transfer is typically less costly than conducting an original valuation study. 

' Can be used as a screening technique to determine if a more detailed, original valuation 

study should be conducted. 

Disadvantages 

' It may be difficult or impossible to find high-quality, well-documented original studies 

from which to obtain unit value estimates that can be appropriately applied to the new 

study site. The use of lower quality unit value estimates will adversely affect the 

accuracy and reliability of the benefit transfer analysis. 

' Unit value estimates can quickly become dated. 

Example Benefits Transfer Study (Bowen et al., 1993; U.S. EPA, 1995) 

In order to estimate the value of recreational fishing in Massachusetts Bays, Bowen et al. 

reviewed several studies of different types of marine recreational fishing experiences, largely 

using the travel cost model. They chose to use estimates reported by Rowe (1985) ranging from 

$13 to $104 (1981 dollars) per fishing day. They then inflated these estimates to 1989 dollars 

($18 to $142) and applied them to the 2.5 million marine recreational fishing trips estimated to 

have been taken in the Massachusetts Bays region in 1989. This calculation yielded a net benefit 

value range of all recreational fishing trips in the Massachusetts Bays of $45 to $355 million. 

This estimate is only reliable as an indication of the order of magnitude of the likely net 

recreational fishing benefits generated by the Bays, because the data on the number of trips 

conducted in the Bays system are subject to considerable uncertainty. In addition, an assumption 

was made that the range of recreational fishing values developed in a variety of different settings 

for a variety of different species reported by Rowe are applicable to the Bays system. The use of 

fishing day values from these other studies to value Massachusetts Bays recreation is also subject 

to criticism because of the use of estimates from a distinctly different geographic region. 

Page 155 
References and Further Reading 

Bingham, T., et al., eds. 1992. Proceedings of the Association of Environmental and Resource 

Economists (AERE) Conference on Benefits Transfers. Washington D.C. 

Bowen, R.E., J.H. Archer, D.G. Terkla, and J.C. Myers. 1993. The Massachusetts Bays 

Management System: A Valuation of Bays Resources and Uses and an Analysis of its 

Regulatory and Management Structure. Boston, Massachusetts: Massachusetts Bays Program. 

Boyle, K.J. and J.C. Bergstrom. 1992. “Benefits Transfer Studies: Myths, Pragmatism, and 

Dealism.” Water Resources Research 28: 657-663. 

Desvousges, W.H., M.C. Naughton, and G.R. Parsons. 1992. “Benefit Transfer: Conceptual 

Problems in Estimating Water Quality Benefits Using Existing Studies.” Water Resources 

Research 28: 675-683. 

Downing, M. and T. Ozuna, Jr. 1994. Testing the Reliability of the Benefit Function Transfer 

Approach. Oak Ridge, Tennessee: Environmental Sciences Division, Oak Ridge Laboratory. 

Kirchhoff, S., B.G. Colby, and J.T. LaFrance. 1997. “Evaluating the Performance of Benefit 

Transfer: An Empirical Inquiry.” Journal of Environmental Economics and Management 33(1): 

75-93. 

Krupnick, A.J. 1993. “Benefits Transfers and Valuation of Environmental Improvements.” 

Resources. 

Loomis, J.B. 1992. “The Evolution of a More Rigorous Approach to Benefit Transfer: Benefit 

Function Transfer.” Water Resources Research 28(3): 701-705. 

Morey, E.R. “What Is Consumer Surplus per Day of Use, When Is it Content Independent of the 

Number of Days of Use, and What Does it Tell Us about Consumer’s Surplus?” Journal of 

Environmental Economics and Management 26: 257-270. 

Opaluch J.J. and M.J. Mazzotta. 1992. “Fundamental Issues in Benefit Transfer and Natural 

Resource Damage Assessment.” in Benefits Transfer: Procedures, Problems, and Research 

Needs. Snowbird, Utah: Workshop Proceedings, Association of Environmental and Resource 

Economists. 

Rowe, R.W. 1985. Valuing Marine Recreational Fishing on the Pacific Coast. La Jolla, 

California: National Marine Fisheries Service, Southwest Fisheries Center. 

Smith, V.K. 1992. “On Separating Defensible Benefit Transfers from Smoke and Mirrors.” 

Water Resources Research 28: 685-694. 

U.S. EPA, Oceans and Coastal Protection Division. 1995. Assessing the Economic Value of 

Estuary Resources and Resource Services in Comprehensive Conservation and Management 

Page 156 
Plan (CCMP) Planning and Implementation, A National Estuary Program Environmental 

Valuation Handbook. Washington, D.C.: U.S. EPA. 

Page 157
7.0 ISSUES AFFECTING THE ECONOMIC VALUATION OF 

ECOLOGICAL BENEFITS 

This section identifies and briefly discusses some additional issues that should be addressed by 

an economic benefit analysis. These issues include: 

• uncertainty and variability; 

• discounting; and 

• equity. 

Please refer to EPA’s (2000) Guidelines for Preparing Economic Analyses and the other 

references provided for a complete treatment of these topics. 

7.1 UNCERTAINTY AND VARIABILITY 

The variability and uncertainty associated with specific estimates is an important consideration 

in a thorough benefits assessment. As EPA’s (2000) Guidelines note, the issue is “not how to 

avoid uncertainty, but how to account for it and present useful conclusions to those making 

policy decisions.” 

Variability and uncertainty are inherent to ecological and economic assessment, stemming from 

multiple potential sources including estimating the effect of the policy, modeling the fate and 

transport of a pollutant (e.g., air modeling), estimating effects, and valuing the effects (or 

changes in the effects). Variability and uncertainty arise from the inherent variation of natural 

processes as well as from limited knowledge about the many relationships between emissions 

and exposures and effects. 

To assess and present uncertainty, EPA (2000) instructs the analyst to: 

• Present outcomes or conclusions based on expected or most plausible values; 

• Provide descriptions of all known key assumptions, biases, and omissions; 

• Perform sensitivity analyses for key assumptions; and 

• Justify the assumptions used in the sensitivity analysis. 

If this initial assessment of uncertainty is not sufficient, then a more sophisticated analysis is 

required. The appropriate approach depends on the objectives of the analysis and the needs of 

the decisionmakers. Uncertainty and variability might be addressed by: 

• Using Monte Carlo analyses or other probabilistic techniques to estimate a probability 

distribution for the output (e.g., benefits); 

Page 158 
• Discussing and/or incorporating expert judgement regarding the potential range of effects 

and/or benefits (e.g., Delphi methods); or 

• Using meta-analysis to combine estimates of inputs (e.g., risks, values) or outputs (e.g., 

benefits estimates) from multiple studies. 

Accounting for uncertainty and variability can provide a more complete characterization of the 

distribution of benefits than point-estimates. Nonetheless, many sources of uncertainty will 

likely remain unquantified. Any remaining omissions, biases, and data gaps should be described 

qualitatively. 

7.2 DISCOUNTING 

When the benefits of an action accrue over time, such as with lagged and/or cumulative effects, 

the role and importance of discounting needs to be considered in the context of the benefits 

assessment. The discount rate used and time period for comparison can have significant effects 

on the magnitude of the benefits estimate and the conclusions of the benefits assessment, 

especially if benefits and costs occur in different points in time. Discounting can be applied to 

monetary values as well as quantitative assessments of benefits. 

Traditionally, present value costs and benefits have been calculated using the shadow price of 

capital or the consumption rate of interest as the discount rate. These may be appropriate or 

inappropriate discount rates, depending on the assumptions made regarding the flow of capital 

and the value of future consequences (e.g., are future values adjusted upward to reflect increased 

value due to increased scarcity). Furthermore, a different discount rate (or even no discounting) 

might be appropriate for inter-generational effects. With respect to the time period of 

comparison, the analysis might choose to translate future values into present ones -- the 

traditional approach -- or alternatively, annualize costs and benefits or accumulate benefits (and 

costs) forward to some future time period. 

Chapter 6 of EPA’s (2000) Guidelines provides a lengthy and detailed discussion of discounting 

as well as numerous references for further reading on this topic. 

7.3 DISTRIBUTIONAL AND EQUITY ANALYSES 

Distributional and equity analyses examine the realized impacts or improvements across 

different sectors of society. Determinations regarding whether a policy or action is “equitable” 

rely on ethical and moral principles, rather than economic principles. In measuring changes in 

social welfare, economists most often implicitly assume that the welfare of all individuals is 

weighted equally. This assumption implies that if a positive change, or benefit, experienced by a 

wealthier individual is determined to be greater in value than the cost, or negative effect, 

experienced by a poorer person, social welfare is said to be “improved” by the change. 

However, such a change may not be “equitable” from an ethical or moral perspective. 

To support a distributional or equity analysis, an ecological benefit analysis should provide 

information on the distribution of costs and benefits (i.e., track who in society is benefitting from 

Page 159 
the change and who is not) in addition to the total or net benefit estimate. The elements of a 

distributional or equity analysis include: 

• Identifying the groups and entities of concern (e.g., race, income) for an equity 

evaluation; 

• Ensuring that data are developed for the groups and entities of concern; and 

• Estimating the distribution of changes across each group and entity of concern. 

Decisionmakers then use the results of the distributional or equity analysis, along with the results 

of the ecological benefit analysis, other analyses conducted, and moral, legal, and/or 

philosophical considerations to evaluate the proposed policy or action. 

Chapter 9 of EPA’s (2000) Guidelines provides a lengthy and detailed discussion of 

distributional analyses as well as numerous references for further reading on this topic. 

References and Further Reading 

Arnold, F.S. 1995. Economic Analysis of Environmental Policy and Regulation. John Wiley 

and Sons, Inc. New York, New York. 

Freeman, A. M. III. 1993. The Measurement of Environmental and Resource Values: Theory 

and Methods. Resources for the Future. Washington, DC. 

Morgan, M.G. and M. Henrion. 1990. Uncertainty: Dealing with Uncertainty in Quantitative 

Risk and Policy Analysis. Cambridge University Press. New York, New York. 

U.S. EPA. 1997. Discounting in Environmental Policy Evaluation, Draft Final Report. 

Prepared by Frank Arnold, Fran Sussman, and Leland Deck for the U.S. EPA, Office of Policy, 

Planning, and Evaluation. April 1, 1997. 

U.S. EPA. 1997. Evaluating the Equity of Environmental Policy Options Based on the 

Distribution of Economic Effects, Draft. Prepared for U.S. EPA, Office of Policy, Planning, and 

Evaluation. May 23, 1997. 

U.S. EPA. 1997. Technical Assistance on a Review and Evaluation of Procedures Used to 

Study Issues of Uncertainty in the Conduct of Economic Cost-Benefit Research and Analysis, 

Draft Report. Prepared by Hagler Bailly Consulting, Incorporated for the U.S. EPA, Office of 

Policy, Planning, and Evaluation. May 27, 1997. 

U.S. EPA. 2000. Guidelines for Preparing Economic Analyses. U.S. EPA, Office of the 

Administrator. EPA/240/R-00/003. September. 

Page 160 
8.0 REFERENCES

8.1 ECOLOGICAL REFERENCES AND FURTHER READING 

Allen, T.F. and Hoekstra, T.W. 1992. Toward a Unified Ecology. Complexity in Ecological 

Systems Series. New York, NY: Columbia University Press. 

Ankley, G.T.; Thomas, N.A.; Di Toro, D.M.; et al. 1994. Assessing potential bioavailability of 

metals in sediments: a proposed approach. Environ. Manage. 18: 331-337. 

Bartell, S.M., R.H. Gardner, and R.V. O'Neill. 1992. Ecological Risk Estimation. Boca Raton, 

FL: Lewis Publishers. 

Beyer, W.N.; Heinz, G.H.; Redmon-Norwood, A.R. (eds.). 1996. Environmental Contaminants 

in Wildlife: Interpreting Tissue Concentrations. A Special Publication of the Society of 

Environmental Toxicology and Chemistry (SETAC), La Point, T.W. (series ed.). Boca Raton, 

FL: CRC Press, Inc., Lewis Publishers. 

Bingham, G., R. Bishop, M. Brody, D. Bromley, E. Clark, W. Cooper, R. Costanza, T. Hale, G. 

Hayden, S. Kellert, R. Norgaard, B. Norton, J. Payne, C. Russell, and G. Suter. 1995. “Issues in 

Ecosystem Valuation: Improving Information for Decision Making.” Ecological Economics 14: 

73-90. 

Briand, F., and Cohen, J.E. 1987. Environmental correlates of food chain length. Science 238: 

956-960. 

Brown, J.H. and Lomolino, M.V. 1998. Biogeography. 2nd Ed. Sunderland, MA: Sinauer 

Associates, Inc. Publishers. 

Cairns, J. Jr.; Niederlehner, B.R. 1995. Ecological Toxicity Testing: Scale, Complexity, and 

Relevance. Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Calabrese, E.J.; Baldwin, L.A. 1993. Performing Ecological Risk Assessments. New York, NY: 

Lewis Publishers. 

Calow, P. (ed.). 1993. Handbook of Ecotoxicology, Volume 1. Boston, MA: Blackwell 

Publishers. 

Cochran, W.G. 1977. Sampling Techniques. 3rd ed. New York, NY: John Wiley and Sons, 

Inc. 

Cochran, W.G.; Cox, G.M. 1957. Experimental Design. New York, NY: Wiley. 

Cockerham, L.G.; Shane, B.S. (eds.). 1994. Basic Environmental Toxicology. Boca Raton, FL: 

CRC Press, Inc., Lewis Publishers. 

Page 161 
Cooke, A.S. 1971. Selective predation by newts on frog tadpoles treated with DDT. Nature 

229: 275-276. 

Costanza, R., R. d’Arge, R. de Groot, S. Farber, M. Grasso, B. Hannon, K. Limburg, S. Naeem, 

R.V. O’Neill, J. Paruelo, R.G. Raskin, P. Sutton, and M. van den Belt. 1988. The value of the 

world’s ecosystem services and natural capital. Ecological Economics 25(1):3-15. 

Costanza, R.; Cumberland, J.; Daly, H.; Goodland, R.; and Norgaard, R. 1997. An Introduction 

to Ecological Economics. Boca Raton, FL: St. Lucie Press. 

Costanza, R.; Low, B.S.; Ostrom, E.; and Wilson, J. (eds.). 2001. Institutions, Ecosystems, and 

Sustainability. Ecological Economics Series. Boca Raton, FL: Lewis Publishers. 

Davis, W.S.; Simon, T.P. 1995. Biological Assessment and Criteria: Tools for Water Resource 

Planning and Decision Making. Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Daily, G.C. (ed). 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. 

Washington, DC: Island Press. 

Daily, G.C., S. Alexander, P.R. Ehrlich, L. Goulder, J. Lubchenco, P.A. Matson, H.A. Mooney, 

S. Postel, S. H. Shneider, D. Tilman, and G.M. Woodwell. 1997. Ecosystem services: benefits 

supplied to human societies by natural ecosystems. Ecological Society of America, Issues in 

Ecology, Number 2, Spring 1997. 

DeBellevue, E.B., T. Maxwell, R. Costanza, and M. Jacobsen. 1993. “Development of a 

Landscape Model for the Patuxent River Watershed.” Discussion Paper #10, Maryland 

International Institute for Ecological Economics, Solomons, MD. 

ESA Ad Hoc Committee on Ecosystem Management. 1995. The Scientific Basis for Ecosystem 

Management. Washington, DC: Ecological Society of America. 

Fitz, H.C., R. Costanza, and E. Reyes. 1993. The Everglades Landscape Model (ELM): 

Summary Report of Task 2, Model Development. Report to the South Florida Water 

Management District, Everglades Research Division. 

Fitz, H.C., E.B. DeBellevue, R. Costanza, R. Boumans, T. Maxwell, L. Wainger, and F. Sklar. 

1995. “Development of a General Ecosystem Model (GEM) for a Range of Scales and 

Ecosystems. Ecological Modeling (in press). 

Foran, J.A. and Ferenc, S.A. (eds.). 1999. Multiple Stressors in Ecological Risk and Impact 

Assessment. A Special Publication of SETAC. Pensacola, FL: Society of Environmental 

Toxicology and Chemistry. 

Freedman, B. 1989. Environmental Ecology. The Impacts of Pollution and Other Stresses on 

Ecosystem Structure and Function. New York, NY: Academic Press. 

Gotelli, N.J. 1998. A Primer of Ecology. 2nd Ed. Sunderland, MA: Sinauer Associates Inc. 

Page 162 
Green, R.H. 1979. Sampling Design and Statistical Methods for Environmental Biologists. 

New York, NY: Wiley. 

Gulland, J.A. 1977. Fish Population Dynamics. London, UK: John Wiley & Sons. 

Hall, S. J.; Raffaelli, D. G. 1993. Food web: theory and reality. In: Begon, M.; Fritter, A. H., 

eds. Advances in Ecological Research, Vol. 24. San Diego, CA: Academic Press; pp. 187-239. 

Hairston, N.G. Jr., Hairston, N.G. Sr. 1993. Cause-effect relationships in energy flow, trophic 

structure, and interspecific interactions. Am. Nat. 142: 379-411. 

Hamelink, J.L.; Landrum, P.F.; Bergman, H.L.; Benson, W.H. (eds). 1994. Bioavailability: 

Physical, Chemical, and Biological Interactions. Boca Raton, FL: CRC Press, Inc., Lewis 

Publishers. 

Hoffman, D.J.; Rattner, B.A.; Burton, G.A. Jr.; Cairns, J., Jr. (eds.). 1995. Handbook of 

Ecotoxicology. Ann Arbor, MI: CRC Press, Inc., Lewis Publishers. 

Hugget, R.J.; Kimerle, R.A.; Mehrle, P.M., Jr.; Bergman, H.L. 1992. Biomarkers: 

Biochemical, Physiological, and Histological Markers of Anthropogenic Stress. A Special 

Publication of the Society of Environmental Toxicology and Chemistry (SETAC), Ward, C.H.; 

Walton, B.T; La Point, T.W. (series eds.). Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Kabata-Pendias, A.; Pendias H. 1984. Trace Elements in Soils and Plants. Boca Raton, FL: 

CRC Press, Inc. 

Kendall, R.J.; Lacher, T.E. (eds.). 1994. Wildlife Toxicology and Population Modeling: 

Integrated Studies of Agroecosystems. A Special Publication of the Society of Environmental 

Toxicology and Chemistry (SETAC), La Point, T.W. (series ed.). Boca Raton, FL: CRC Press, 

Inc., Lewis Publishers. 

Klemm, D.J.; Lewis, P.A.; Fulk, F.; Lazorchak, J.M. 1990. Macroinvertebrate Field and 

Laboratory Methods for Evaluating the Biological Integrity of Surface Waters. Washington, 

DC: U.S. EPA. EPA/600/4-90/030. 

Krebs, C.J. 1978. Ecology: The Experimental Analysis of Distribution and Abundance; Second 

Edition. New York, NY: Harper & Row. 

Landis, W.G.; Yu, M. 1995. Introduction to Environmental Toxicology: Impacts of Chemicals 

upon Ecological Systems. Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Landis, W.G.; Hughes, J.S.; Lewis, M.A. (eds.). 1993. Environmental Toxicity and Risk 

Assessment. Philadelphia, PA: American Society for Testing and Materials. 

MacArthur, R.H., and Wilson, E.O. 1963. An equilibrium theory of insular biogeography. 

Evolution 17: 373-387. 

Page 163 
MacArthur, R.H., and Wilson, E.O. 1967. The Theory of Island Biogeography. Princeton, NJ: 

Princeton University Press. 

Margalef, R. 1968. Perspectives in Ecological Theory. Chicago, IL: University of Chicago 

Press. 

Maughan, J.T. 1993. Ecological Assessment of Hazardous Waste Sites. New York, NY: Van 

Nostrand Reinhold. 

Moltmann, J.F.; Römbke, J. 1996. Applied Ecotoxicology. Boca Raton, FL: CRC Press, Inc., 

Lewis Publishers. 

National Research Council (NRC). 1994. Science and Judgment in Risk Assessment. 

Washington, DC: National Academy Press. 

National Research Council (NRC). 1993. Issues in Risk Assessment. Washington, DC: 

National Academy Press. 

National Research Council (NRC). 1983. Risk Assessment in the Federal Government: 

Managing the Process. Washington, DC: National Academy Press. 

Neilson, A.H. 1994. Organic Chemicals in the Aquatic Environment: Distribution, 

Persistence, and Toxicity. Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Newman, M.C. 1995. Quantitative Methods in Aquatic Ecotoxicology. Boca Raton, FL: CRC 

Press, Inc., Lewis Publishers. 

Newman, M.C.; McIntosh, A.W. (eds.). 1991. Metal Ecotoxicology: Concepts and 

Applications. Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Norse, E. 1990. Threats to Biological Diversity in the United States. Report prepared for the 

U.S. EPA, Washington, DC, by Industrial Economics, Contract No. 68-W8-0038, Work 

Assignment 115. 

Noss, R.P.; LaRoe, E.T.; ans Scott, J.M. 1995. Endangered Ecosystems of the United States: A 

Preliminary Assessment of Loss and Degradation. Washington, DC: U.S. Department of the 

Interior, National Biological Service. 

Odum, E.P., in collaboration with H.T.Odum. 1959. Fundamentals of Ecology. Philadelphia, 

PA: Saunders. 

Ostrander, G. (ed.). 1996. Handbook of Aquatic Toxicology Methods. Boca Raton, FL: CRC 

Press, Inc., Lewis Publishers. 

Ott, W.R. 1995. Environmental Statistics and Data Analysis. Boca Raton, FL: CRC Press, 

Inc., Lewis Publishers. 

Page 164 
Parker, S.P. (ed.). 1994. Dictionary of Scientific and Technical Terms; Fifth Edition. New

York, NY: McGraw-Hill.

Pimentel, D. 1988. Economic benefits of natural biota. Ecological Economics 25(1):45-47.

Pimm, S.L. 1980. Properties of food webs. Ecology 61: 219-225.

Pimm, S.L. 1982. Food Webs. New York, NY: Chapman and Hall

Principe, P.P. 1995. Ecological benefits assessment: A policy-oriented alternative to regional

ecological risk assessment. Human and Ecological Risk Assessment 1(4):423-435.

Rand, G.M.; Petrocelli, S.R. 1985. Fundamentals of Aquatic Toxicology. Methods and 

Applications. New York, NY: McGraw Hill.

Renzoni, A.; Fossi, M.C.; Lari, L.; Mattei, N. (eds.). 1994. Contaminants in the Environment. 

A Multidisciplinary Assessment of Risks to Man and Other Organisms. Boca Raton, FL: CRC

Press, Inc., Lewis Publishers.

Ricklefs, R.E. 1990. Ecology; Second Edition. New York, NY: W.H. Freeman.

Scodari, P. 1992. Wetland Protection Benefits. Draft Report. Prepared for U.S. EPA, Office of

Policy, Planning, and Evaluation under Grant No. CR-817553-01. October.

Stephan, C.E., Mount, D.I., Hanson, D.J., Gentile, J.H., Chapman, G.A., and Brungs. W.A. 

1985. Guidelines for Deriving Numeric National Water Quality Criteria for the Protection of 

Aquatic Organisms and Their Uses. Duluth, Minnesota: U.S. EPA. NTIS No. PB85-227049. 

Sullivan, T.F. 1993. Environmental Regulatory Glossary. Government Institutes, Inc. 

Suter, G.W. II. 1993. Ecological Risk Assessment. Boca Raton, FL: Lewis Publishers. 

Suter, G.W. II. 1989. “Ecological Endpoints.” in Warren-Hicks, W., B.R. Parkhurst, and S.S. 

Baker, Jr., eds. Ecological Assessment of Hazardous Waste Sites: A Field and Laboratory 

Reference Document. EPA Document 600/3-89/013. Corvallis Environmental Research 

Laboratory, Oregon. 

Suter, G. W., Efroymson, R.A., Sample, B.E., Jones, D.S. 2000. Ecological Risk Assessment for 

Contaminated Sites. Boca Raton, FL: Lewis Publishers. 

Terborgh, J. 1989. Where Have All the Birds Gone? Princeton, NJ: Princeton University Press. 

Trapp, S.; McFarlane, J.C. (eds.). 1995. Plant Contamination: Modeling and Simulation of 

Organic Chemical Processes. Boca Raton, FL: CRC Press, Inc., Lewis Publishers. 

Page 165 
U.S. Department of the Interior (U.S. DOI). 1987. Guidance on Use of Habitat Evaluation 

Procedures and Suitability Index Models for CERCLA Application. Washington, DC: U.S. Fish 

and Wildlife Service, National Ecology Center; PB86-100151. 

U.S. EPA. 1983. Environmental Effects of Regulatory Concern Under TSCA: A Position 

Paper. Washington, DC: Office of Toxic Substances, Health and Environmental Review 

Division (Author: Clements, R.G.) 

U.S. EPA. 1986. Guidelines for Deriving Numerical Criteria for the Protection of Aquatic 

Organisms and Their Uses. Washington, DC: Office of Water Regulations and Standards. 

U.S. EPA. 1986. Guidelines for the Health Risk Assessment of Chemical Mixtures. 

Washington, DC: Office of Health and Environmental Assessment; EPA/600/8-87/045. 

U.S. EPA. 1986. Quality Criteria for Water 1986. Washington, DC: Office of Water 

Regulations and Standards; EPA/440/5-86/001. 

U.S. EPA. 1988. Superfund Exposure Assessment Manual. Washington, DC: Office of Solid 

Waste and Emergency Response Directive 9285.5-1; EPA/540/1-88/001. 

U.S. EPA. 1989. Ecological Assessment of Hazardous Waste Sites: A Field and Laboratory 

Reference. Corvallis, OR: Office of Research and Development, Environmental Research 

Laboratory; EPA/600/3-89/013. 

U.S. EPA. 1989. Rapid Bioassessment Protocols for Use in Streams and Rivers: Benthic 

Macroinvertebrates and Fish. Washington, DC: Office of Water; EPA/444/4-89/001 (Authors: 

Plafkin, J.L.; Barbour, M.T.; Porter, K.D.; Gross, S.K.; Hughes, R.M.). 

U.S. EPA. 1989. Risk Assessment Guidance for Superfund: Volume 2 - Environmental 

Evaluation Manual, Interim Final. Washington, DC: Office of Solid Waste and Emergency 

Response; EPA/540/1-89/001A. 

U.S. EPA. 1989. Scoping Study of the Effects of Soil Contamination on Terrestrial Biota. 

Washington, DC: Office of Toxic Substances. 

U.S. EPA. 1989. Superfund Exposure Assessment Manual -- Technical Appendix: Exposure 

Analysis of Ecological Receptors. Athens, GA: Office of Research and Development, 

Environmental Research Laboratory (December). 

U.S. EPA. 1990. Biological Criteria, National Program Guidance for Surface Waters. 

Washington, DC: Office of Water Regulations and Standards; EPA/440/5-90/004. 

U.S. EPA. 1990. Ecosystem Services and Their Valuation. Prepared by RCG/Hagler, Bailly, 

Inc., for the Office of Policy, Planning, and Evaluation. Washington, DC: U.S. EPA. 

U.S. EPA. 1990. Managing Contaminated Sediments: EPA Decision-Making Processes. 

Washington, DC: Sediment Oversight Technical Committee; EPA/506/6-90/002. 

Page 166 
U.S. EPA. 1990. National guidance: wetlands and nonpoint source control programs. 

Memorandum from Martha G. Prothro, Director, Office of Water Regulations and Standards; 

Washington, DC: Office of Water (June 18). 

U.S. EPA. 1991. Assessment and Control of Bioconcentratable Contaminants in Surface 

Waters. June 1989 Draft prepared by U.S. EPA's National Effluent Toxicity Assessment Center, 

Environmental Research Laboratory - Duluth, MN. Washington, DC: Office of Water 

Regulations and Standards; and Cincinnati, OH: Office of Health Effects Assessment. 

U.S. EPA. 1991. Ecological Exposure and Effects of Airborne Toxic Chemicals: An Overview. 

Corvallis, OR: Office of Research and Development, Environmental Research Laboratory; 

EPA/600/3-91/001. 

U.S. EPA. 1991. Summary Report on Issues in Ecological Risk Assessment. Washington, DC: 

Risk Assessment Forum; EPA/625/3-91/018. 

U.S. EPA. 1991. Technical Support Document for Water Quality-based Toxics Control. 

Washington, DC: Office of Water; EPA/505/2-90/001. 

U.S. EPA. 1991. The Watershed Protection Approach Framework Document. Office of 

Wetlands, Oceans, and Watersheds. Washington, DC: U.S. EPA. 

U.S. EPA. 1992. Dermal Exposure - Principles and Applications; Final; Washington, DC: 

Office of Health and Environmental Assessment; EPA/600/8-91/011B. 

U.S. EPA. 1992. Developing a Work Scope for Ecological Assessments. ECO Update, 

Intermittent Bulletin, Volume 1, Number 4. Washington, DC: Office of Emergency and 

Remedial Response, Hazardous Site Evaluation Division; Publ. 9345.0-05I. 

U.S. EPA. 1992. Framework for Ecological Risk Assessment. Washington, DC: Risk 

Assessment Forum; EPA/630/R-92/001. 

U.S. EPA. 1992. Guidance on Risk Characterization for Risk Managers and Risk Assessors. 

February 26 Memorandum from F. Henry Habicht II, Deputy Administrator, to U.S. EPA 

Assistant Administrators and Regional Administrators. Washington, DC: Office of the Deputy 

Administrator. 

U.S. EPA. 1992. Guidelines for Exposure Assessment. Federal Register. 57: 22888-22938 

(May 29). 

U.S. EPA. 1992. Peer Review Workshop Report on a Framework for Ecological Risk 

Assessment. Washington, DC: Risk Assessment Forum; EPA/625/3-91/022. 

U.S. EPA. 1992. Report on the Ecological Risk Assessment Guidelines Strategic Planning 

Workshop. Washington, DC: Risk Assessment Forum; EPA/630/R-92/002. 

Page 167 
U.S. EPA. 1992. Science Advisory Board's Review of the Draft Final Exposure Assessment 

Guidelines (SAB Final Review Draft, August 1991). Washington, DC: Science Advisory 

Board; EPA/SAB/IAQC-92/015. 

U.S. EPA, Office of Policy Planning and Evaluation. 1992. Biological Populations as 

Indicators of Environmental Change. Washington, DC: U.S. EPA. EPA/230/R-92/011 

U.S. EPA. 1993. A Guidebook to Comparing Risks and Setting Environmental Priorities. 

Washington, DC: U.S. EPA. EPA/230/B-98/003. 

U.S. EPA. 1993. A Review of Ecological Assessment Case Studies from a Risk Assessment 

Perspective. Washington, DC: Risk Assessment Forum; EPA/630/R-92/005. 

U.S. EPA. 1993. Guidance for Planning for Data Collection in Support of Environmental 

Decision Making Using the Data Quality Objectives Process. Interim Final. Quality Assurance 

Management Staff; EPA QA/G-4. 

U.S. EPA. 1993. Guidance for Specifying Management Measures for Sources of Nonpoint 

Pollution in Coastal Waters. Office of Water. Washington, DC: U.S. EPA. EPA/840/B-92/002. 

U.S. EPA. 1993. Habitat Evaluation: Guidance for the Review of Environmental Impact 

Assessment Documents. Prepared by Dynamac Corporation for the Office of Federal Activities 

under EPA Contract No. 68-C0-0070. January. 

U.S. EPA. 1993. Wildlife Exposure Factors Handbook Volumes I and II. Washington, DC: 

Office of Research and Development; EPA/600/R-93/187ab. 

U.S. EPA. 1994. A Review of Ecological Assessment Case Studies from a Risk Assessment 

Perspective, Volume II. Washington, DC: Office of Research and Development, Risk 

Assessment Forum; EPA/630/R-94/003. 

U.S. EPA. 1994. Background for NEPA Reviewers: Grazing on Federal Lands. Prepared by 

Science Applications International Corporation under EPA Contract No. 68-C8-0066. February. 

U.S. EPA. 1994. Ecological Risk Assessment Issue Papers. Washington, DC: Office of 

Research and Development, Risk Assessment Forum; EPA/630/R-94/009. 

U.S. EPA. 1994. Establishing Background Levels. Quick Reference Fact Sheet. Washington, 

DC: Office of Solid Waste and Emergency Response. OSWER Directive 9285.7-19FS. 

Publication PB94-963313; EPA/540/F-94/030. 

U.S. EPA. 1994. Guidance for the Data Quality Objectives Process; EPA QA/G-4. 

Washington, DC: Quality Assurance Management Staff; Final, September. 

U.S. EPA. 1994. Managing Ecological Risks at EPA: Issues and Recommendations for 

Progress. Prepared by M.E. Troyer and M.S. Brody. Washington, DC: U.S. EPA. EPA/600/R94/183. 

Page 168 
U.S. EPA. 1994. Memorandum from Carol Browner, Administrator, to Assistant 

Administrators concerning "Toward a Place-Driven Approach: The Edgewater Consensus on an 

EPA Strategy for Ecosystem Protection. May 24. 

U.S. EPA. 1994. Peer Review Workshop Report on Ecological Risk Assessment Issue Papers. 

Washington, DC: Office of Research and Development, Risk Assessment Forum; EPA/630/R94/008. 

U.S. EPA. 1994. Selecting and Using Reference Information in Superfund Ecological Risk 

Assessments. ECO Update, Intermittent Bulletin, Volume 2, Number 4. Washington, DC: 

Office of Emergency and Remedial Response, Hazardous Site Evaluation Division; Publication 

9345.10I; EPA/540/F-94/050; NTIS PB94-963319. 

U.S. EPA. 1994. Toward a Place-Driven Approach: The Edgewater Consensus on an EPA 

Strategy for Ecosystem Protection. Ecosystem Protection Workgroup. Washington, DC: U.S. 

EPA. March 15 Draft. 

U.S. EPA. 1995. Draft Science Policy Council Statement on EPA Policy: Cumulative Risk 

Framework, With a Focus on Improved Characterization of Risks for Multiple Endpoints, 

Pathways, Sources, and Stressors. Washington, DC: Science Policy Council. 

U.S. EPA. 1995. Ecological Risk: A Primer for Risk Managers. Washington, DC: U.S. EPA. 

EPA/734/R-95/001. 

U.S. EPA. 1995. Ecological Significance and Selection of Candidate Assessment Endpoints. 

ECO Update, Intermittent Bulletin, Volume 3, Number 1. Washington, DC: Office of 

Emergency and Remedial Response, Hazardous Site Evaluation Division; Publication 9345.0-

11FSI; EPA/540/F-95/037; NTIS PB95-963323. 

U.S. EPA. 1995. EPA Risk Characterization Policy. March 21 Memorandum from Carol 

Browner, Administrator, to EPA staff. Washington, DC: Office of the Administrator. 

U.S. EPA. 1995. Great Lakes Water Quality Initiative Technical Support Document for the 

Procedure to Determine Bioaccumulation Factors. Washington, DC: Office of Water, U.S. 

EPA. EPA/820/B-95/005. 

U.S. EPA. 1997. Guidance on Cumulative Risk Assessment. Washington, DC: Science Policy 

Council. 

U.S. EPA. 1997. Priorities for Ecological Protection: An Initial List and Discussion Document 

for EPA. EPA/600/S-97/002. Washington, DC: U.S. EPA. 

U.S. EPA. 1998. Guidelines for Ecological Risk Assessment. EPA/630/R-95/002B. 

Washington, DC: U.S. EPA. 

Page 169 
U.S. EPA. 2000. Assessing the Neglected Ecological Benefits of Watershed Management 

Practices: A Resource Book. Prepared for the Assessment and Watershed Protection Division, 

Office of Water by ICF Consulting. Washington, DC: U.S. EPA. April. 

U.S. EPA. 2001. Risk Characterization Handbook. Memorandum from W. Michael McCabe, 

Deputy Administrator, Office of the Administrator; Washington, DC: Office of the 

Administrator (January 10). 

Wilcove, D.S. 1985. Nest predation in forest tracts and the decline of migratory songbirds. 

Ecology 66: 1211-1214. 

Wilson, E.O. and Bossert, W.H. 1971. A Primer of Population Biology. Stamford, CT: Sinauer 

Associates Inc. Publishers. 

Page 170 
8.2 ECONOMIC REFERENCES AND FURTHER READING 

Adams, R.M. and T.D. Crocker. 1991. "Materials Damages," in Braden, John B. and Charles D. 

Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. North-Holland, 

Amsterdam: Elsevier Science Publishers. 

Ahearn, M.C. 1997. “Why Economists Should Talk to Scientists and What They Should Ask: 

Discussion.” Journal of Agricultural and Applied Economics, July, 29(1): 113-116. 

Alberini, A. 1995. “Optimal Designs for Discrete Choice Contingent Valuation Surveys: 

Single-Bounded, Double-Bounded, and Bivariate Models.” Journal of Environmental 

Economics and Management, 28(3): 287-306. 

Arnold, F.S. 1995. Economic Analysis of Environmental Policy and Regulation. John Wiley 

and Sons, Inc. New York, New York. 

Arrow, K., R. Solow, P.R. Portney, E.E. Leamer, R. Radner, and H. Schuman. 1993. “Report of 

the NOAA Panel on Contingent Valuation.” Federal Register, January 15, 58(10): 4601-4614. 

Bateman, I.J. and K.G. Willis, Eds. 1998. Valuing Environmental Preferences: Theory and 

Practice of the Contingent Valuation Method in the U.S., E.U., and Developing Countries. 

Oxford University Press, Oxford. 

Bartik, T.J. 1988. “Evaluating the Benefits of Non-marginal Reductions in Pollution Using 

Information on Defensive Expenditures.” Journal of Environmental Economics and 

Management, 15: 111-127. 

Bartik, T.J. 1988. “Measuring the Benefits of Amenity Improvements in Hedonic Price 

Models.” Land Economics 64(2): 172-183. 

Bayless, M. 1982. “Measuring the Benefits of Air Quality Improvements: A Hedonic Salary 

Approach.” Journal of Environmental Economics and Management 9(2): 81-99. 

Bertollo, P. 1998. “Assessing Ecosystem Health in Governed Landscapes: A Framework for 

Developing Core Indicators.” Ecosystem Health 4(1): 33-51. 

Bingham, T., et al., eds. 1992. Proceedings of the Association of Environmental and Resource 

Economists (AERE) Conference and Benefits Transfers. Washington, D.C. 

Bishop, R.C. and T.A. Heberlein. 1979. “Measuring Values of Extramarket Goods: Are 

Indirect Measures Biased?” American Journal of Agricultural Economics, December: 926-929. 

Bjornstad, D.J. and J.R. Kahn, eds. 1996. The Contingent Valuation of Environmental 

Resources: Methodological Issues and Research Needs. Brookfield, Vermont: Edgar Elgar 

Publishing Ltd. 

Page 171 
Bockstael, N.E., K.E. McConnell, and I.E. Strand. 1991. “Recreation” in Braden, John B. and 

Charles D. Kolstad, eds. Measuring the Demand for Environmental Quality. North-Holland, 

Amsterdam: Elsevier Science Publishers. 

Bockstael, N.E., K.E. McConnell, and I. Strand. 1989. “A Random Utility Model for 

Sportfishing: Some Preliminary Results for Florida.” Marine Resource Economics 6: 245-260. 

Bockstael, N.E., K.E. McConnell, and I.E. Strand. 1989. “Measuring the Benefits of 

Improvements in Water Quality: The Chesapeake Bay.” Marine Resource Economics 6(1): 1­

18. 

Bowen, R.E., J.H. Archer, D.G. Terkla, and J.C. Myers. 1993. The Massachusetts Bays 

Management System: A Valuation of Bays Resources and Uses and an Analysis of its 

Regulatory and Management Structure. Boston, Massachusetts: Massachusetts Bays Program. 

Boyle, K.J. and J.C. Bergstrom. 1992. “Benefits Transfer Studies: Myths, Pragmatism, and 

Dealism.” Water Resources Research 28: 657-663. 

Braden, J.B. and C.D. Kolstad, eds. 1991. Measuring the Demand for Environmental Quality. 

North-Holland, Amsterdam: Elsevier Science Publishers. 

Cameron, T.A. 1992. “Combining Contingent Valuation and Travel Cost Data for the Valuation 

of Nonmarket Goods.” Land Economics, August, 68(3): 302-17. 

Carson, R.T. et al. 1996. “Was the NOAA Panel Correct About Contingent Valuation?” 

Washington, D.C.: Resources for the Future. 

Cole, R.A., et al. 1996. Linkages Between Environmental Outputs and Human Services, IWR 

Report 96-R-4. Prepared for U.S. Army Corps of Engineers, Evaluation of Environmental 

Investment Research Program. 

Cooper, J.C. 1993. “Optimal Bid Selection for Dichotomous Choice Contingent Valuation.” 

Journal of Environmental Economics and Management, 24(1): 25-40. 

Cooper, J.C. and J. Loomis. 1992. “Sensitivity of Willingness to Pay to Bid Design in 

Dichotomous Choice Contingent Valuation.” Land Economics, 68(2): 211-224. 

Costanza, R. et al. 1997. “The Value of the World’s Ecosystem Services and Natural Capital.” 

Nature 387: 253-260. May. 

Cummings, R.C., D. S. Brookshire, and W.D. Schulze, Eds. Valuing Environmental Goods: An 

Assessment of the Contingent Valuation Method. Rowan and Allanheld Publishers, Totowa, NJ. 

Daily, G., ed. 1997. Nature’s Services: Societal Dependence on Natural Ecosystems. 

Washington, D.C.: Island Press. 

Page 172 
Desvousges, W.H., M.C. Naughton, and G.R. Parsons. 1992. “Benefit Transfer: Conceptual 

Problems in Estimating Water Quality Benefits Using Existing Studies.” Water Resources 

Research 28: 675-683. 

Desvousges, W., V.K. Smith, and M.P. McGivney. 1983. A Comparison of Alternative 

Approaches for Estimating Recreation and Related Benefits of Water Quality Improvement. 

Prepared for the U. S. Environmental Protection Agency. Report 230-05-83-001. Washington 

D.C. 

Downing, M. and T. Ozuna, Jr. 1994. Testing the Reliability of the Benefit Function Transfer 

Approach. Oak Ridge, Tennessee: Environmental Sciences Division, Oak Ridge Laboratory. 

Edwards, S.F. and G.D. Anderson. 1984. “Land Use Conflicts in Coastal Zone: An Approach 

for the Analysis of the Opportunity Costs of Protecting Coastal Resources.” Journal of the 

Northeastern Agricultural Economics Council 13(1): 78-81. 

Englin, J.E., T.A. Cameron, R.E. Mendelsohn, G.A. Parsons, and S.A. Shankle. 1991. 

Valuation of Damages to Recreational Trout Fishing in the Upper Northeast due to Acidic 

Deposition. Richland, Washington: Pacific Northwest Laboratory. Prepared for National Acidic 

Precipitation Assessment Program. 

Ehrlich, P. and A. Ehrlich. 1997. Betrayal of Science and Reason. Island Press, Washington, 

D.C. 

Fletcher, J., W. Adamowicz, and T. Graham-Tomasi. 1990. “The Travel Cost Model of

Recreation Demand.” Leisure Sciences 12: 119-147.

Freeman, A.M., III. 1993. The Measurement of Environmental and Resource Values: Theory

and Methods. Washington, D.C.: Resources for the Future.

Gramlich, Edward M. 1990. A Guide to Benefit-Cost Analysis, Second Edition. Prentice Hall.

Haab, T.C. and K.E. McConnell. “Referendum Models and Negative Willingness to Pay:

Alternative Solutions.” Journal of Environmental Economics and Management, 32(3): 251-270. 

Hanemann, W.M. 1991 “Willingness to Pay and Willingness to Accept: How Much Can They

Differ?” American Economic Review 81(3): 635-647.

Hanemann, W.M. 1984. “Welfare Evaluations in Contingent Valuation Experiments with

Discrete Responses.” American Journal of Agricultural Economics, August, 66: 332-341. 

Hanley, N. and C.L. Spash. 1993. Cost Benefit Analysis and the Environment. Brookfield,

Vermont: Edward Elgar Publishing Limited.

Hanley, N., J.F. Shogren, and B.White. 1997. Environmental Economics in Theory and

Practice. Oxford University Press, New York, New York. p. 362-364, 395-396. 

Page 173 
Johnson, F.R., W.H. Desvousges, L.L. Wood, and E.E. Fries. 1995. Conjoint Analysis of 

Individual and Aggregate Environmental Preferences, Technical Paper No. T-9502. Triangle 

Economic Research. 

Just, R.E., D.L. Hueth, and A. Schmitz. 1982. Applied Welfare Economics and Public Policy. 

Englewood Cliffs, New Jersey: Prentice-Hall. 

Kannimen, B.J. 1993. “Design of Sequential Experiments for Contingent Valuation Studies.” 

Journal of Environmental Economics and Management, 25(1): s1-s11.

 Kaoru, Y., V.K. Smith, and J.L. Liu. 1995. “Using Random Utility Models to Estimate the 

Recreational Value of Estuarine Resources.” American Journal of Agricultural Economics, 

February, 77: 141-151. 

King, D.M. 1997. Using Ecosystem Assessment Methods in Natural Resource Damage 

Assessment, Paper #2. Prepared for U.S. Department of Commerce, NOAA, Damage 

Assessment and Restoration Program. 

Kirchhoff, S., B.G. Colby, and J.T. LaFrance. 1997. “Evaluating the Performance of Benefit 

Transfer: An Empirical Inquiry.” Journal of Environmental Economics and Management 33(1): 

75-93. 

Krupnick, A.J. 1993. “Benefits Transfers and Valuation of Environmental Improvements.” 

Resources. 

Loomis, J.B. 1993. Integrated Public Lands Management: Principles and Applications to 

National Forests, Parks, Wildlife Refuges, and BLM Land. New York, New York: Columbia 

University Press. 

Loomis, J.B. 1992. “The Evolution of a More Rigorous Approach to Benefit Transfer: Benefit 

Function Transfer.” Water Resources Research. 28(3): 701-705. 

Loomis, J., A. Gonzales-Caban, and R. Gregory. 1994. “Do Reminders of Substitutes and 

Budget Constraints Influence Contingent Valuation Estimates?” Land Economics. 70(4): 499­

506. 

Mackenzie, J. 1992. “Evaluating Recreation Trip Attributes and Travel Time via Conjoint 

Analysis.” Journal of Leisure Research 24(2): 171-184. National Recreation and Park 

Association. 

McConnell, K. and I. Strand. 1981. “Measuring the Cost of Time in Recreation Demand 

Analysis.” American Journal of Agricultural Economics: 153-156. 

Milon, J.W., C. Kiker, and D. Lee. 1997. “Ecosystem Management and the Florida Everglades: 

The Role of Social Scientists.” Journal of Agricultural and Applied Economics, July, 29(1): 99­

107. 

Page 174 
Mitchell, R.C. and R.T. Carson. 1989. Using Surveys to Value Public Goods: The Contingent 

Valuation Method. Washington, D.C.: Resources for the Future. 

Mitchell, R.C. and R.T. Carson. 1986. The Use of Contingent Valuation Data for Benefit/Cost 

Analysis in Water Pollution Control. Washington, D.C.: Resources for the Future. 

Mitchell, R.C. and R.T. Carson. 1984. An Experiment in Determining Willingness to Pay for 

National Water Quality Improvements. Washington, D.C.: Resources for the Future. 

Morey, E.R. “What Is Consumer Surplus per Day of Use, When Is it Content Independent of the

Number of Days of Use, and What Does it Tell Us about Consumer’s Surplus?” Journal of

Environmental Economics and Management 26: 257-270.

Morgan, M.G. and M. Henrion. 1990. Uncertainty: Dealing with Uncertainty in Quantitative

Risk and Policy Analysis. Cambridge University Press. New York, New York.

Musser, W.N. 1997. “Why Economists Should Talk to Scientists and What They Should Ask:

Discussion.” Journal of Agricultural and Applied Economics, July, 29(1): 109-112.

Opaluch J.J. and M.J. Mazzotta. 1992. “Fundamental Issues in Benefit Transfer and Natural

Resource Damage Assessment.” in Benefits Transfer: Procedures, Problems, and Research

Needs. Snowbird, UT: Workshop Proceedings, Association of Environmental and Resource

Economists.

Palmquist, R. 1991. “Hedonic Methods.” in Braden, John B. and Charles D. Kolstad, eds. 

1991. Measuring the Demand for Environmental Quality. North-Holland, Amsterdam: Elsevier

Science Publishers.

Palmquist, R.B., F.M. Fritz, and T. Vukina. 1997. “Hog Operations, Environmental Effects, and

Residential Property Values.” Land Economics 73(1): 114-124.

Pearce, David W. 1998. “Auditing the Earth.” Environment. 40(2): 23-28.

Pearce, David W. 1993. Economic Values and the Natural World. The MIT Press, Cambridge,

Massachusetts. 

Pearce, D.W. and R.K. Turner. 1990. Economics of Natural Resources and the Environment.

Maryland: The Johns Hopkins University Press.

Pimentel, D., C. Wilson, C. McCullum, R. Huang, P. Dwen, J. Flack, Q. Tran, T. Saltman, B.

Cliff. 1997. “Economic and Environmental Benefits of Biodiversity.” BioScience. 47(11):

747-757.

Principe, P. 1995. “Ecological Benefits Assessment: A Policy-Oriented Alternative to Regional

Ecological Risk Assessment.” Human and Ecological Risk Assessment 1(4): 423-435.

Randall, A., B. Ives, and C. Eastman. 1974. “Bidding Games for Valuation of Aesthetic

Environmental Improvements.” Journal of Environmental Economics 1: 132-149.

Page 175 
Rowe, R.W. 1985. Valuing Marine Recreational Fishing on the Pacific Coast. La Jolla, 

California: National Marine Fisheries Service, Southwest Fisheries Center. 

Scodari, P. 1992. Wetland Protection Benefits. Draft Report. Prepared for the Office of Policy, 

Planning, and Evaluation, U.S. EPA. Grant No. CR-817553-01. 

Smith, V.K. 1992. “On Separating Defensible Benefit Transfers from Smoke and Mirrors.” 

Water Resources Research 28: 685-694. 

Smith, V. K. 1989 “Taking Stock of Progress with Travel Cost Recreation Demand Methods: 

Theory and Implementation.” Marine Resource Economics 6: 279-310. 

Tietenberg, T. 1992. Environmental and Natural Resource Economics. Harper Collins 

Publisher. 

U.S. EPA. 2000. Guidelines for Preparing Economic Analyses. U.S. EPA, Office of the 

Administrator. EPA/240/R-00/003. September. 

U.S. EPA. 1997. A Conceptual Model for the Economic Valuation of Ecosystem 

Damages Resulting from Ozone Exposure. Draft Report. Prepared by Science Applications 

International Corporation, for the U.S. EPA, Office of Air Quality Planning and Standards. 

U.S. EPA. 1997. Discounting in Environmental Policy Evaluation, Draft Final Report, 

Prepared by Frank Arnold, Fran Sussman, and Leland Deck for the U.S. EPA, Office of Policy, 

Planning, and Evaluation. April 1, 1997. 

U.S. EPA. 1997. Evaluating the Equity of Environmental Policy Options Based on the 

Distribution of Economic Effects, Draft. Prepared for U.S. EPA, Office of Policy, Planning, and 

Evaluation. May 23, 1997. 

U.S. EPA. 1997. Technical Assistance on a Review and Evaluation of Procedures Used to 

Study Issues of Uncertainty in the Conduct of Economic Cost-Benefit Research and Analysis, 

Draft Report. Prepared by Hagler Bailly Consulting, Incorporated for the U.S. EPA, Office of 

Policy, Planning, and Evaluation. May 27, 1997. 

U.S. EPA, Oceans and Coastal Protection Division. 1995. Assessing the Economic Value of 

Estuary Resources and Resource Services in Comprehensive Conservation and Management 

Plan (CCMP) Planning and Implementation, A National Estuary Program Environmental 

Valuation Handbook. Washington, D.C.: U.S. EPA. 

U.S. EPA, RTI. 1983. A Comparison of Alternative Approaches for Estimating Recreation and 

Related Benefits of Water Quality Improvements, EPA Document 230-05-83-001 Under 

Cooperative Agreement #68-01-5838. Washington D.C.: U.S. EPA. 

Viscusi, W. K. 1993. “The Value of Risks to Life and Health.” Journal of Economic 

Literature. XXXI(4): 1912-1946. 

Page 176 
Whittington, D., et al. 1994. The Economic Value of Improving the Environmental Quality of 

Galveston Bay. Galveston National Estuary Program. Publication GBNEP-3B. 

Willig, R. 1976. “Consumer Surplus Without Apology.” American Economic Review 66(4): 

589-597. 

Willis, K. and G. Garrod. 1991. “An Individual Travel Cost Method of Evaluating Forest 

Recreation.” Journal of Agricultural Economics 42(1): 33-42. 

Page 177