10019983
UNITED STATES
ENVIRONMENTAL PROTECTION AGENCY
FRAMEWORK FOR THE
ECONOMIC ASSESSMENT or
ECOLOGICAL BENEFITS
DRAFT
U.S. EPA Headquarters Library
Mail code 3201
1200 Pennsylvania Avenue NW
Washington DC 20460
EPA
1007
1998.3
July 1998
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FRAMEWORK FOR THE ECONOMIC
ASSESSMENT OF ECOLOGICAL BENEFITS
DRAFT
Prepared for
Ecological Benefit Assessment Workgroup
Social Sciences Discussion Group
Science Policy Council
U.S. Environmental Protection Agency
Prepared by
U.S. EPA Headquarters Library
ICF Incorporated Mail code 3201
9300 Lee Highway 12°0 Pennsylvania Avenue NW
Fairfax, VA 22031 W^lngton DC 20460
July 1998
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NOTICE
This is a draft EPA document intended for internal review. It is not to be released outside the
Agency and should not be cited or quoted. Questions about or comments on this document
should be sent to John D. Harris (5204 G), 703-603-9075, fax 703-603-9104
Keys to using this document:
• This document provides a conceptual framework designed to promote increased
communication between natural scientists conducting an ecological assessment and
policy analysts conducting an economic benefit analysis. The goal is to improve the
economic assessment of the ecological benefits of maintaining, protecting or restoring
ecological resources. The document establishes an analytical structure, gives relevant
information on methodologies, provides references to other sources of information, and
notes results from other studies. It is not a "cookbook" that gives detailed instructions on
all aspects of conducting an ecological assessment or an economic benefit analysis.
• This document is designed as a tool for economists and policy analysts who are charged
with evaluating the benefits of actions that affect ecological systems. However, it should
be of use to many different audiences, such as ecologists supporting economic analyses,
risk assessors or decision makers.
• Because economists and policy analysts having some familiarity with economics are the
primary audience, terms are used with meanings typical in the economics field, which
may vary substantially from meanings used by ecologists or others. A primary example is
the title phrase "ecological benefits," which an economist might interpret as relating to the
identification and valuation of benefits associated with ecological systems or services.
An ecologist might interpret the phrase as relating to something that improves the
structure and function of the ecological system. Ultimately, the document will contain a
glossary to help minimize misinterpretation of terms.
• This document is designed to be used electronically, and, in electronic form, will have
hypertext links between sections, a searchable bibliography database, and a searchable
database of economic valuation studies. The intent is to make the final version of the
document available through the EPA Home Page on the Internet. It can, of course, also
be used in hard copy, but this format loses the linking and search features. A
WordPerfect hard copy version will be available to download from the EPA Internet site.
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TABLE OF CONTENTS
1.0 Introduction and Overview 1
1.1 Introduction 1
1.2 Overview of This Document 1
2.0 Interdisciplinary Coordination 4
2.1 Introduction 4
2.2 Identifying Endpoints for the Economic Benefit Analysis and Linkages Between
the Ecological Changes and the Economic Benefit Endpoints 5
2.3 Conducting a Qualitative and/or Quantitative Assessment of Effects on the
Economic Benefit Endpoints and Selecting Specific Endpoints for a Monetized
Analysis of Economic Benefits 10
2.4 Ensuring Analytical and Data Compatibility 12
2.5 Additional Issues 14
3.0 Important Principles of Ecology and Ecological Assessment 17
3.1 Defining an Ecosystem and Other Levels of Ecological Organization 17
3.2 Understanding the Interactive Nature of an Ecosystem 18
3.3 Measuring/Assessing Impacts to Ecological Components 21
4.0 The Ecological Risk Assessment Process 26
4.1 Overview of Ecological Risk Assessment 26
4.2 Phase I: Problem Formulation 30
4.3 Phase II: Analysis Phase 39
4.4 Phase III: Risk Characterization 51
5.0 Background Theory on Valuing Changes to Ecological Resources 55
5.1 Defining the Economic Value of an Ecological Resource 55
5.2 Measuring the Benefits of Improvements to Ecological Resources —
The Concept of Willingness-to-Pay 55
5.3 Measuring the Benefits of Improvements to Ecological Resources —
Theoretical Basis 57
5.4 Measuring the Benefits of Improvements to Ecological Resources —
Estimating Willingness-to-Pay 59
6.0 Economic Assessment of Ecological Benefits 61
6.1 Components of an Economic Assessment of Ecological Benefits 61
6.2 Identifying the Service Flows and Other Values Provided by an Ecological
Resource 64
6.2.1 Direct, Market Uses 67
6.2.2 Direct Non-Market Uses 69
6.2.3 Indirect Non-Market Uses 71
6.2.4 Non-Market, Non-Use Values 73
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6.3 Approaches to Measuring Resource Values 75
6.3.1 Market-Price and Supply/Demand Relationships 80
6.3.2 Market-Based Valuation Approaches 83
6.3.3 Travel Cost Methodologies 89
6.3.4 Random Utility Model 93
6.3.5 Hedonic Price and Hedonic Wage Methodologies 96
6.3.6 Contingent Valuation 102
6.3.7 Combining Contingent Valuation with Other Approaches: Contingent
Activity 110
6.3.8 Conjoint Analysis and Contingent Ranking 112
6.3.9 Benefits Transfer 117
7.0 Issues Affecting the Economic Valuation of Ecological Benefits 122
7.1 Uncertainty and Variability 122
7.2 Aggregation 123
7.3 Discounting 123
7.4 Distributional and Equity Analyses 124
8.0 References and Data Bases 126
8.1 Ecological References and Further Reading 126
8.2 Economic References and Further Reading 128
8.3 Illustrative Sample of Data Sources 134
8.4 Data Base of Ecological Resource Values 136
8.5 Bibliographic Data Base 137
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1.0 INTRODUCTION AND OVERVIEW
l.l INTRODUCTION
The Social Sciences Discussion Group (SSDG), convened under the auspices of the EPA Science
Policy Council, is working to improve the Agency's ability to conduct economic analyses. 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 to a common approach for analyzing ecological benefits and a
better understanding of both the scientific and economic techniques used in these analyses. This
framework document is intended to serve both of these purposes.
Representing a joint effort of both ecologists and economists, this framework:
• Summarizes the objectives and processes of an ecological assessment;
• Discusses the elements of an economic benefit analysis and the techniques used to
estimate the economic value of ecological benefits; and
• Identifies the major opportunities for improving the economic benefit analysis
through increased interdisciplinary coordination.
This document is intended to provide general information for EPA staff and others who are
interested in the concepts and techniques used to assess and quantify ecological effects and to
monetize ecological costs and benefits.
The document is not designed to be either a "cookbook" or a "how to" manual — it does not
provide detailed, 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
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).
1.2 OVERVIEW OF THIS DOCUMENT
This framework document for assessing and economically valuing ecological benefits is intended
to serve four general purposes:
• Promoting greater coordination between ecologists and economists working on
such efforts;
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• Providing an understanding of the approaches currently in use;
• Suggesting additional sources of information; and
• Providing a starting point for those individuals who need to study and
economically value the ecological effects of a specific action.
Exhibit 1 presents the conceptual framework for assessing the economic value of ecological
benefits described in this guidance. This framework is intended to provide a starting point for
approaching such analyses; it does not prescribe a particular method of research or interaction.
Exhibit 1
Conceptual Framework for the Economic Assessment of Ecological Benefits
Identify and Characterize
Economic Benefit Endpoints
Direct, Market
Uses
Direct, Non-
Market Uses
Define the Linkages
Between Ecological
Changes and Economic
Benefit Endpoints
Indirect non-
Market Uses
Ecological Assessments
Define Changes to Each
Economic Benefit Endpoint
Conduct a Qualitative
and/or Quantitative
Assessment of Effects
on Economic Benefit
Endpoints
Other Scientific Input
Select Economic Benefit
Endpoints for Monetized
Assessment of Changes
Ensure Analytical
Compatibility
Select and Apply Appropriate
Technique(s) for Valuing Changes
to Selected Benefit Endpoints
Present Results of the Qualitative, Quantitative,
and Monetized Assessment of the Economic
Benefits of Ecological Changes
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This conceptual framework consists of three components:
• Interdisciplinary Coordination (Section 2): Establish the baseline and
scenarios to be examined, establish the connections between the ecological
changes and economic benefit endpoints, identify and define the foci of the
economic benefit analysis, choose a qualitative, quantitative or monetized
assessment of the effects, ensure analytical compatibility between the ecological
assessment and the economic analysis, and ensure the smooth transfer of
information between the ecological assessment and the economic analysis.
• Ecological Assessment and Scientific Input (Sections 3 and 4): Identify and
estimate the ecological changes associated with an action.
• Economic Benefit Analysis (Sections 5 and 6): Identify and characterize the
affected economic benefit endpoints, describe and/or model how the ecological
changes affect each economic benefit endpoint, select a qualitative, quantitative or
monetized assessment of the economic effects, estimate the changes to the
economic benefit endpoints; and select and apply appropriate techniques for
estimating the monetary value of the estimated changes to selected economic
benefit endpoints.
The organization of this document is consistent with these three components. Because the
economic assessment of ecological benefits is not a linear process, there are frequent references
to other sections of the document and electronic links to those sections where additional
information is presented relating to a specific topic or issue.
In addition to the main body of the document, which covers the areas described above, there are
three supplementary sections:
• Issues (Section 7): A discussion of some additional issues relevant to the
economic analysis of ecological benefits, including uncertainty, discounting,
aggregation, and equity.
• Economic Valuation Study Database (Section 8.4): A key-word searchable
database containing information on almost 300 economic valuation studies. The
information provided includes the resource or economic endpoint valued, the
principal investigators, the research methodology, and the resulting value
estimate.
• Bibliographic Database (Section 8.5): A key-word searchable database
containing full bibliographic information for all references mentioned in the
document, as well as additional studies, related journal articles, and reports.
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2.0 INTERDISCIPLINARY COORDINATION
Interdisciplinary coordination is the fundamental element of the conceptual approach laid out in
this document. This section first discusses how coordination between natural scientists (e.g.,
biologists, ecologists, chemists) and economists can improve the economic assessment of
ecological benefits. This section then highlights specific opportunities for coordination and
identifies differences in perspective between the disciplines, which must be recognized for
successful coordination to occur.
2.1 INTRODUCTION
Interdisciplinary coordination refers to the discussion of goals between individuals in different
disciplines and the cooperative efforts of these individuals to achieve these goals. Although the
level of interaction between the disciplines increases, the distinct frameworks of each discipline
are maintained.
Interdisciplinary coordination promotes:
• Better understanding of the structure and function of ecological resources by
economists;
• Better understanding of economic benefit analysis by ecologists;
• Improved definition of economic benefit endpoints; and
• Streamlined data collection.
Successful interdisciplinary coordination requires a basic understanding of the objectives and
approach of other disciplines. Improving this basic understanding is, in fact, one of the main
objectives of this document.
The need for coordination between the disciplines is not restricted to a particular activity, but
rather, factors into each step of the design and implementation of the ecological assessment and
the analysis of economic benefits. There are three specific areas where interdisciplinary
coordination is particularly important to improve the economic assessment of ecological benefits:
D Identifying the endpoints for the economic benefit analysis and understanding the
linkages between the ecological changes and the effects on the economic benefit
endpoints;
D Conducting a qualitative and/or quantitative assessment of the effects on the
economic benefit endpoints and selecting specific endpoints for the monetized
assessment of economic benefits; and
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CD Ensuring that the ecological assessment and the economic benefit analysis are
analytically compatible.
The opportunities for, and benefits of, coordination in each of these areas are discussed in the
following sections.
2.2 IDENTIFYING ENDPOINTS FOR THE ECONOMIC BENEFIT ANALYSIS AND
LINKAGES BETWEEN THE ECOLOGICAL CHANGES AND THE ECONOMIC
BENEFIT ENDPOINTS
The selection of endpoints determines how the
ecological assessment and the economic benefit
analysis will evaluate the effects of the action or
change under study. The endpoints selected for
the ecological assessment may be the same or
different from the benefit endpoints examined
by the economic benefit analysis. In either case,
it is important to understand the linkages
between the ecological assessment endpoints
and the economic benefit endpoints.
The economic benefit analysis is based on the
premise that actions affecting the state of an
ecological resource, measured in terms of
changes to the ecological assessment endpoints,
will result in changes to the goods and services
provided 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, estimating the magnitude and significance of those effects, and developing belter
methods for measuring and modeling the linkages between changes to ecological resources and
changes to the economic benefit endpoints. In addition, by working with economists to define
the economic benefit endpoints, ecologists can help ensure that 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 targeted to the needs
of the economist.
The following two subsections describe two stages of potential improvement in the
identification, definition, and measurement of these linkages:
• Better identification and definition of the linkages between ecological changes
and effects on economic benefit endpoints; and
What Is An Endpoint?
Endpoints differ by discipline. Economic
benefit endpoints are those goods or
services, provided or supported by the
ecological resource, that are valued by
humans either directly or indirectly.
Ecological assessment endpoints are the
ecological components directly or
indirectly affected by the change or action
being examined. Changes to the
ecological assessment endpoints are used
to measure the ecological impact to the
ecosystem due to the action under study.
Changes in the economic benefit
endpoints are used to estimate the
economic value of the action under study.
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• Improved measurement and modeling of these linkages in estimating and valuing
changes to economic benefit endpoints.
Identification and Definition of Linkages
The thoroughness of the economic benefit analysis depends on the identification and definition
of the linkages between changes to the ecological resource(s) and changes to the economic
benefit endpoints. The identification and definition of the linkages between changes to an
ecological resource and changes to specific economic benefit endpoints begins with a qualitative
understanding of the relationships and interactions that occur within the natural system.
Ecologists and other scientists can help the economist to understand these relationships and more
completely define the list of economic benefit endpoints likely to be affected by a change. One
approach is to work together to extend the conceptual model developed as part of the ecological
assessment to identify the economic benefit endpoints that are likely to be affected and the
pathways by which these effects will be realized. Exhibit 2 provides an example of how a
conceptual model might be expanded to include the affected economic benefit endpoints.
Exhibit 2
Expanding a Conceptual Model to Include Linkages to
Specific Economic Benefit Endpoints
Im proved Local Septic System s
I
Reduced N utrient Loading
I
Reduced Eutrophication in Local W aters
Improved Aquatic Habitat
Im proved W etland
Function and Structure
I
Increased
F ish /S h ellfish
Populations
Im proved W a te r Filtration
Improved Storm
P ro te c tio n
Increased
M igratory Bird
Visitation Rate
Increased
Shore Bird
Population
Increased
R ecreational
F ish /S h ellfish
L a n d in g s
I m proved
Bird Watching
O pportunities
I m proved
W a te r
Q u a 1 ity
1m proved
R ecreational
Swimming and
B oating
O pportunities
Reduced
Property Losses
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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 economic benefit endpoints.
A second approach that might be used to more completely identify the affected economic benefit
endpoints is a simple table listing the expected ecological changes and the corresponding effect
of each change on various economic benefit endpoints. Exhibit 3 illustrates this approach by
identifying some possible affected economic benefit endpoints for two ecological changes.
Exhibits
Hypothetical List Linking Ecological Changes to Potential Economic Effects
Ecological Change
Reduced turbidity of water body
Economic Effects
Increased commercial and recreational fish harvests
Reduced water treatment costs
Improved aesthetic quality of the water
Increased wetland acreage
Reduced costs of storm damage
Improved recovery after storm induced combined sewer
overflows
Reduced water treatment costs
Increased commercial and recreational fishery and shellfish
harvest
Increased aesthetic value due to increased diversity or
abundance of nature
Increased property values
Several researchers have explored alternative approaches for better identifying and describing the
linkages between ecological changes and economic effects. King (1997) used a table format,
similar to Exhibit 3, to link specific human uses and values such as hunting, bird watching,
property protection, and water-based recreation, to specific wetland functions including
providing habitat, floodwater storage, and pollutant uptake. The U.S. Army Corps of Engineers
(USAGE) (Cole, et al., 1996) uses similar, although slightly more complex, approaches to
identify and describe the links between the ecological effects of a potential USAGE
environmental restoration project and the likely impacts on the human uses and services
associated with the affected ecological resource(s). For example, one USAGE table on the
effects of building barriers and dredging and filling activities identifies the potential ecological
changes, such as change in area available for use (in acres or square km) or change in shore
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protection capacity (shore miles or acres affected). These changes are linked to those human
uses or services that might be impacted either directly or indirectly, such as recreational boating
or swimming or protection of property. The USAGE tables also provide possible measures of
the ecological change and approaches for measuring and valuing the associated change to the
human uses (i.e., the economic change).
Measuring and Modeling Linkages
Building on a better qualitative understanding of the linkages between ecological changes and
effects on economic benefit endpoints, efforts can expand to better measure and model these
linkages. This section discusses how economic valuation methods currently use ecological data
to estimate the benefits of specific ecological changes, and identifies opportunities for
developing better approaches for measuring and incorporating such linkages in economic benefit
analyses.
Current Practice
Economists often use the experience and data from other disciplines in their analyses. In
estimating the economic value of ecological benefits, most economic valuation models require
some measure of the ecological change (e.g., change in water quality, acres of open space, type
offish living in a water body).
The "measures" of the ecological change used in economic valuation models vary significantly in
level of specificity and complexity. Most often, economic analyses use only a simple measure of
the ecological change, such as a subjective ranking of quality or the measured concentration of a
pollutant in the air or water. Furthermore, most economic valuation studies examine the
economic impact of only one effect (e.g., lost recreational fishing opportunities associated with a
change in water quality), rather than considering the full range of possible impacts of an
ecological change on multiple economic endpoints. Exhibit 4 presents some examples of recent
economic benefit valuation analyses.
Future Opportunities
Better methods are needed for measuring and modeling the connections between ecological
changes and effects on economic benefit endpoints. One goal is to move beyond a relatively
linear approach in which the valuation model focuses on a single change and ignores the
feedback loops and interactions inherent in the natural system. Developing economic models
that link more directly with the ecological assessment model represents one possible opportunity
for improving economic benefit analysis. Because of their ability to adjust to different ecological
conditions, such "linked" valuation models are likely to be applicable in a wider variety of
settings, to be able to be scaled up or down to reflect various levels of ecological change, and to
better capture the effects of interactions inherent in the natural system.
Many ecological assessment models are now capable of modeling complex interactions and
consequences within the natural system:
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Coastal Ecological Landscape Spatial Simulation. The first computer model to
compile a cellular spatial process ecosystem model. This model is capable of
"simulating successional changes from one ecosystem type to another as a
response to human impacts and natural variation" (Fitz, H.C., et al, 1993).
General Ecosystem Model. A computer model capable of recording ecosystem
changes, such as nutrient, light, temperature, and water changes and the
effects/responses by macrophyte and algal communities. It is constructed
modularly, thus allowing easy shifts in application, such as increasing/decreasing
the covered area or changing the community/ecosystem (Fitz, H.C., et al., 1995).
Patuxent Landscape Model. A regional computer model designed to "simulate
environmental impacts generated from changes in land use patterns and practices
in the watershed." It uses software that combines several other software packages
and databases, including the GEM, Spatial Modeling Program, and GIS databases
to calculate how these changes affect nearby cell areas (i.e., land, air, water)
(DeBellevue, E.B., et al, 1993).
Exhibit 4
Examples of Recent Economic Benefit Analyses Incorporating Scientific Data
(exhibit to be expanded in future draft to include EPA studies as examples)
Study Title
Description
Using Random Utility
Models to Estimate the
Recreational Value of
Estuarine Resources.
Estimated a household production function, using a random utility
model framework, to test the effects of nonpoint source pollution on
recreational fishers site choice. The nonpoint source pollutant (nitrogen
and phosphorous) contributions included in the study were estimated
from National Oceanic and Atmospheric Administration's National
Coastal Pollutant Discharge Inventory (NCPDI). Using the NCPDI, the
study also estimated biochemical oxygen demand (BOD). These data
were all included in estimation procedures and found to be statistically
significant (Kaoru, Y., et al, 1995).
Measuring the Benefits
of Improvements in
Water Quality: The
Chesapeake Bay.
Contingent valuation data estimated willingness-to-pay to improve the
Chesapeake Bay's water quality using econometric choice models. The
Bay's water quality was represented by the product of nitrogen and
phosphorous readings, because of the collinearity of the two variables,
from 1977, because that was the last year when complete data were
available. The study concluded that significant willingness-to-pay
exists to improve the Bay's water quality. It also indicated that the
dollar range ($10-100 million) may ultimately be too low as human-use
patterns change to reflect improved water quality variable (Bockstael,
N. E., etal, 1989).
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Similarly complex economic models, including general equilibrium models which account for
feedbacks and interactions through the market system, also exist. In addition, some researchers
have recently defined economic valuation models that explicitly link the modeling of ecological
changes to the modeling of effects on specific benefit endpoints (see Exhibit 5). Clearly, the
capabilities exist for developing more complex and "connected" economic valuation models.
Long-term research goals include the development of models that simultaneously (1) account for
many different ecological functions, and (2) estimate the effects on many different economic
benefit endpoints.
Exhibit 5
A Conceptual Model for the Economic Valuation of Ecosystem
Damages Resulting from Ozone Exposure
A conceptual model that links ecological and economic models is under development to address the
complex interactions among ecosystem components and the behavior of economic agents. The
conceptual model is applied in a case study to demonstrate the feasibility of application, the
possibilities for ecosystem value measurement, and the limitations imposed by the data requirements.
The case study examines the impact of changes in ambient ozone concentrations on a southern pine
ecosystem. An ecological model, TREGRO, is used to estimate the impact of elevated ozone
concentrations on measures of tree growth. These results are combined with other relationships to
demonstrate how such ozone impacts might indirectly affect the status of other ecosystem
components, specifically the nesting habitat for red-cockaded woodpeckers. Economic benefits are
realized through outdoor activities, such as hunting and wildlife watching, provided as services of
the southern pine forest ecosystem. The model uses a biodynamic growth equation to capture the
relationship between economic activity and the status of the ecosystem. The case study is limited
geographically, by the ecosystem components considered, by the stressors applied, and by the
techniques to apply the ecological impacts to an expanded domain. In addition, this conceptual
framework is augmented by including feedback effects from the economic system to the ecosystem
(EPA, 1997).
Interdisciplinary research efforts, as discussed in this section, require a better understanding of
what researchers in other fields are doing and a willingness to collaborate. By providing
information on the process and objectives of the ecological assessment and discussing
opportunities for improved cooperation and coordination between scientists and economists, this
document hopes to encourage continued discussions and interest in expanding the connections
between the ecological assessment and the economic benefit analysis.
2.3 CONDUCTING A QUALITATIVE AND/OR QUANTITATIVE ASSESSMENT OF
EFFECTS ON THE ECONOMIC BENEFIT ENDPOINTS AND SELECTING
SPECIFIC ENDPOINTS FOR A MONETIZED ANALYSIS OF ECONOMIC
BENEFITS
In addition to identifying and defining the economic benefit endpoints, economists and ecologists
should how the economic benefit analysis will measure and describe the changes to these
endpoints. The economist's decisions should be informed by the ecologist. The goal of the
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economic benefit analysis is to capture as much of the potential benefits of an action in the
monetized assessment of benefits. For many reasons, as discussed below, it may not be possible
or practical to estimate the monetary value of changes for all the economic benefit endpoints.
Thus, the economist must determine how and at what depth the economic benefit analysis will
examine the effects on each of the economic benefit endpoints. In determining whether a
qualitative, quantitative, or monetized analysis is appropriate for each economic benefit endpoint,
the discussion should center on:
Q Time and resource constraints;
Q Analytical purpose;
Q Data availability;
Q Anticipated magnitude of the impact to the benefit endpoint relative to the
anticipated impacts on other endpoints;
Q Anticipated uncertainty of the impact to the benefit endpoint, relative to
uncertainty regarding other endpoints;
Q Compatibility with the data and outputs of the ecological assessment; and
Q Availability of appropriate measurement techniques.
Time and resource constraints generally require that the economic benefit analysis focus on
quantifying and monetizing the benefits associated with only a few benefit endpoints. The
objective is to select those benefits endpoints that are expected to experience the greatest
economic benefits. Effects on other benefit endpoints, which are not monetized, should be
assessed quantitatively or qualitatively. Economists should work with ecologists and other
scientists to determine which benefit endpoints are likely to experience the greatest change. For
example, what economists might consider relatively minor ecological changes could have
widespread and/or long-term consequences that result in a significant economic effect.
The appropriate type of assessment of the economic benefits is often determined by answering
the questions: "Why is the analysis being conducted?" and "What are the questions which the
analysis will address?" In some cases, a qualitative or quantitative assessment of the economic
value of the ecological benefits, rather than a monetized assessment, may be all that is necessary
to support the decision-making process.
Often decisions regarding how the effects on specific benefit endpoints are analyzed and
described are based on the availability of appropriate data techniques to support a quantitative or
monetized assessment of the changes. Some ecological services may be too complex and too
poorly understood to quantify or monetize potential changes (e.g., carbon cycle, nitrogen cycle).
When appropriate data is not available to support a monetized assessment of the economic
benefits, the economic benefit analysis should still include a qualitative or quantitative
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assessment of the effect.
A thorough economic benefit assessment focuses not just on the effects that can be monetized,
but on the full scope of effects. Qualitative, and when possible quantitative, measures of effects
on economic benefit endpoints are necessary because they provide a measure of a service's
importance and the degree of change, even when a dollar value cannot be assigned to that
change. Furthermore, a thorough qualitative and quantitative assessment of the changes can
supplement and support the monetize assessment of changes.
2.4 ENSURING ANALYTICAL AND DATA COMPATIBILITY
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. As a
result, it is important that the baseline from which the effects are measured and the specifics of
the action or change (i.e., the scenario) that is considered are consistent between the ecological
assessment and the economic benefit analysis. In defining the scenario(s) to be examined, it is
important to specifically define the action or change to be evaluated, the area(s) expected to be
affected, the time period under study, and what other factors or actions (e.g., other regulations)
might affect the outcome and how they will be accounted for in the ecological assessment and
the economic benefit analysis.
A common understanding of the baseline from which effects are to be measured and the specifics
of the scenario to be analyzed is the first step toward ensuring that:
• Outputs from the ecological assessment are compatible with the needs of the
economic benefit analysis; and
• Findings from the ecological assessment and the conclusions of the economic
benefit analysis are analytically consistent.
Additional time and effort is often required, however, to determine the best approach for meeting
the data needs of the economic benefit analysis. Economists need to understand the data
traditionally collected by an ecological assessment and determine how well these data can
address the data needs of the economic analysis. Ecologists also need to better understand the
data needs of the economic benefit analysis. Economists need to work with ecologists from the
initiation of the ecological assessment, if possible, to develop better methods for translating and
transferring data from the ecological assessment to the economic benefit analysis.
What can be and is measured in the ecological assessment will dictate, in part, what ecological
changes and economic effects are captured by the economic benefit analysis. It is important to
understand exactly what changes can be identified, how those changes will be described or
measured (e.g., at what level of spatial or temporal detail, with what level of certainty), and
acknowledge the major limitations or uncertainties associated with those descriptions or
measurements.
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Spatial and Temporal Distribution of Data. The spatial and temporal
distribution of effects is of particular concern in the transfer of information. The
magnitude of the economic value of an action affecting an ecological resource
may depend on the spatial and temporal distribution of the effects. Therefore, to
accurately estimate the potential ecological effects and appropriately value the
change to the ecological resource, the ecological assessment and economic benefit
analysis must analyze a consistent spatial and temporal distribution of effects. In
addition, the ecological assessment and economic analysis should be consistent in
accounting for variation in the intensity and frequency of the event under
examination and the time path of recovery if the resource is damaged.
An ecological assessment might focus its examination of the effects of an action
at the point where the exposure and effects are the greatest and determine the
acceptability of those point effects (e.g. local or individual organism level effects).
The economic benefit analysis, however, should consider the full range of effects
and assess the variability of those effects across the whole area and population of
concern. It is also important for the economic benefit analysis to consider the
potential impacts to populations, ecosystems, or regions, which may have a large
spatial scale.
In addition, when ecological changes are assessed as if they occurred
instantaneously following some action, it is important for the purposes of
assessing the economic value of the ecological changes that the effects be
assessed across the entire time period under examination. Furthermore, 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, and determining an appropriate
discounting scheme for comparing changes in future effects against changes in
current effects (see Issues section for further discussion on discounting).
Data Limitations and Uncertainty. The economic benefits analysis should
attempt to account for the uncertainties surrounding the ecological assessment
process as well as the imprecision of most economic valuation techniques. The
level of uncertainty in the ecological assessment process as well as the economic
valuation process is sometimes substantial.
Although many economists acknowledge the imprecision of the available
economic valuation techniques, they often do not specifically account for this
uncertainty in presenting the value estimate. Economists also do not typically
account for the uncertainty in the ecological assessment when they develop value
estimates. Although the results of an ecological assessment, for example, are
often expressed as a probability with associated uncertainties, there is no
standardized methodology for utilizing this type of measure in an economic
benefit analysis, (see Issues section for further discussion on accounting for
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variability and uncertainty.)
Through improved coordination, economists and scientists can work together to determine how
to address the data needs of the economic benefits analysis and to solve issues regarding the
analytical compatibility of the ecological assessment and economic benefits analysis. For
example, if the baseline from which changes are measured and the scenarios examined are
consistent between the ecological assessment and economic benefit analysis, there are likely to
be fewer difficulties in ensuring that the data provided by the ecological assessment and the
design of the economic analysis are compatible.
2.5 ADDITIONAL ISSUES
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. This section identifies some of the 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.
Further discussion of this issue can be found in the sections on Important
Principles of Ecology and Ecological Assessment and Background on
Economic Theory.
•• 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 includes the disciplines agreeing on the units of measurement
as well as analytical scope and termination points (e.g., an occurrence with local
versus widespread ecological and economic effects).
• Focus. The definition of critical endpoints, and the change(s) of interest, likely
differ for each discipline. How each discipline proposes to examine the changes
may differ (i.e., economics tends to measure change linearly, without system or
feedback analyses). These differences are partially a consequence of each
discipline's training but may also reflect reality (e.g., an important ecological
effect may not be economically measurable).
• Metrics. Economists typically want to standardize effects or welfare changes into
dollars in order to compare effects that may be dissimilar. It should be recognized
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that other metrics may also be appropriate for qualitative and quantitative
descriptions of ecological and economic effects.
References and Further Reading
Aheam, 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.
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).
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.
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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.
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. 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.
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3.0 IMPORTANT PRINCIPLES OF ECOLOGY AND
ECOLOGICAL ASSESSMENT
This section defines some basic terms and concepts used by ecologists and explains how these
concepts can be applied in conducting an ecological assessment. First, this section defines an
ecosystem and levels of ecological organization. Secondly, the interactions that occur within
ecosystems are described, including the concepts of "food chain," "food web," and "energy
flow." Finally, this section briefly introduces the various types of ecological assessments.
3.1 DEFINING AN ECOSYSTEM AND OTHER LEVELS OF ECOLOGICAL
ORGANIZATION
Ecology is the study of biological systems in which there are interactions between living
organisms and their environment — these systems are called ecosystems. The concept of an
ecosystem can be applied at any scale ranging, for example, from a small pond to an entire
mountain 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."
There are five primary levels of ecological organization:
• Individual,
• Population,
• Community,
• Ecosystem, and
• Landscape.
A species is a group of individuals that are able to interbreed. In a species, slight biological
variations will exist among individuals. A population is a group of organisms of the same
species that occupies a particular space over a given interval of time. A community is an
organized assemblage or association of populations in a prescribed area or a specific habitat. An
ecosystem, as described in more detail below, can be viewed as a biotic (living) community
functioning within its abiotic (non-living) environment. A group of ecosystems make up a
landscape. A landscape may be composed of several isolated or interactive ecosystems and are
usually defined in geographic terms, such as a prairie or mountain.
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 or energy flow. An important feature is that the ecosystems function as units with
nutrients and energy flowing between the different biotic and abiotic components.
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In an ecosystem, community and habitat 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. Organisms themselves can also
affect the physical environment through their reactions, thereby indirectly impacting other
organisms.
Interactions occur among individuals within a population, between individuals of different
populations, and between organisms and their physical surroundings. 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 populations. An example 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. As a result of these complex interactions, the consequences on individual
plants or animals are usually slight compared with the combined effects on the community or
ecosystem level. Therefore, most ecosystems are constantly shining and changing, never existing
within definitive boundaries or operating in isolation.
3.2 UNDERSTANDING THE INTERACTIVE NATURE OF AN ECOSYSTEM
Ecosystems may be as large as unbroken tracts of forest and grassland or smaller than a pond.
The ecosystem is an energy-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 (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 system; and
« Dead organic material and inorganic substrate that act as short-term nutrient pools
and support the cycling of nutrients within the system.
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 the final dissipation of energy.
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The food chain 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 is 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 chain (see Exhibit 6).
A simple food chain might be: oak leaf & caterpillar & small bird o hawk. One useful approach
in defining food chains is to group organisms based on their trophic levels (i.e., their position in
the food chain).
Exhibit 6
Quaternary consumers
(carnivores)
Tertiary consumers
(carnivores)
Secondary consumers
(carnivores)
Primary consumers
(herbivores)
Producers
Trophic Level Organization
The major categories for trophic level organization are producers, primary consumers, and
secondary consumers. However, ecosystems are too complex to be characterized by a single,
unbranched food chain. Instead, food chains are usually interlinked to form a food web (see
Exhibit 7).
A food web is a trophic system formed by a series of interconnecting food chains.
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Exhibit 7
Terrestrial Food Web
Grasses and seeds
Bushes with fruit
C83006-2
The food web for most communities is very complex, involving hundreds or thousands of
organisms. 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 instances, 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.
Two processes occur in an ecosystem through the food web: energy 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 and nutrient cycles through the individual organisms, the ecologist is
able to analyze the changes in an entire ecosystem.
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Example of Translation and Magnification of the Effects
of Pollutant Discharges Through the Food Web
The effects of pollutant discharges can influence not only the health, behavior and survival of
individuals, 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 the population
and community. In the same way that energy flows from one trophic level to the next, certain
pollutants can also be transferred and bioaccumulate as they travel up the food web. For
instance, pollutants entering rivers, lakes, estuaries and other water bodies can bioaccumulate
to very high levels in biotic tissues. These levels may become toxic and adversely affect
individuals, populations, communities and whole ecosystems, despite the fact that these
pollutants were originally introduced at relatively low levels. For example, mercury is
generally present in small amounts in surface waters and is consumed by fish as well as taken
up by algae. However, as the algae and fish are consumed by higher organisms, the mercury
accumulates in these organism and is not excreted. Therefore, organisms at higher trophic
levels may have mercury concentrations in their tissues, which far exceeds the concentration
of mercury in the surrounding water.
3.3 MEASURING/ASSESSING IMPACTS TO ECOLOGICAL COMPONENTS
Ecological assessment is the 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 predator
species), chemical (e.g., presence of a toxic chemical), and physical (e.g., loss of habitat). Each
of these factors must be included in assessing the impacts on an ecosystem component. Thus,
understanding how various factors of an ecosystem interact is critical to assessing the effect of a
change to any component(s) of the ecosystem: The effect(s) of a change must be traced through
the food web in order to understand the full magnitude of the change to the ecosystem.
There are many types of ecological assessments that provide a basis or support information for an
economic benefit analysis. Several assessments, such as a hazard assessment or a habitat
assessment, may be included under another more inclusive assessment, such as an ecological risk
assessment. Although every type of ecological assessment is not completely unique, each one
displays considerable divergence in the way problems are formulated and analyzed. This section
briefly describes the different types of ecological assessments that may be performed.
Ecological Risk Assessment
An ecological risk assessment is a process used to predict and/or evaluate the effects of human
activities or a natural phenomena on an ecological resource(s). Ecological effects may be
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evaluated both qualitatively or quantitatively in terms of changes to structural components,
functional components and levels of organization. EPA recently issued draft final guidelines for
conducting an ecological risk assessment (EPA, 1998). At EPA, ecological risk assessments are
among the most common type of ecological assessment. Other types of ecological assessments,
such as those conducted under the National Environmental Policy Act (NEPA), are usually
similar yet not completely the same.
In terms of ecological risk assessment, 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 change is called a "stressor." 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 (i.e., energy
flows through the food web). Assessment endpoints are the ecological components that the
assessment examines.
Environmental Assessment (EA)
An EA is frequently required under 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. Preparation of an EA requires a scoping effort to identify environmental
areas that may be impacted by the proposed action. Contacting and coordinating efforts with
appropriate agencies (e.g., the Fish and Wildlife Service for the potential presence of threatened
and endangered species, the Army Corps of Engineers for wetlands, etc.) is also necessary when
conducting an EA. If significant environmental impacts are identified, the EA must provide a
full discussion of the impacts, direct and indirect, including the impacts of alternative
actions/uses. There must also be an evaluation of the probable cumulative, long-term
environmental effects including any beneficial impacts.
Environmental Impact Statement (EIS)
An El S is a type of assessment that attempts to reveal the consequences of a proposed action as
an aid to governmental decision-making. In the U.S., 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. An EIS can vary significantly in its scope and content due to the
variety of activities that are considered "major federal actions". Also, a large number of
personnel are required from diverse disciplines, including ecologists, biologists, social scientists
and environmental engineers. An EIS is required to predict any or all future effects on the
environment. As a result, an EIS 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 other types of assessments.
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Habitat Assessment
Habitat assessments evaluate the suitability of an ecosystem's habitat for a species. This
involves predicting biological, physical, and chemical effects that may impair the habitat.
Habitat assessments provide a broad context for estimating the effects of pollution and providing
a means for resource managers to incorporate these effects into their management models. The
most conspicuous 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 (Scodari, 1992). 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.,
recreational versus navigational), the extent to which pollution is impacting these uses, and the
necessary pollution control measures that are needed. Attainability analyses must consider
habitat limitations such as frequency of low tides, water quality, and physical structure of the
habitat.
Models are commonly used to elaborate habitat assessments. For instance, the
Environmental Requirements and Pollution Tolerance System (ERAPT) is a large
relational database on habitat requirements of aquatic species in the upper midwestera
U.S. that allows inferences to be made about water quality from particular species
present. The model is able to make this inference, because some species are more
sensitive to changes in water quality and pollution levels than others.
Hazard Assessment
Hazard assessments determine the existence of a hazard. This type of assessment predicts the
effects of chemical, biological, or physical stressors on the environment by extrapolating effects
observed in the laboratory and comparing them to those expected in the field. In risk
assessments that try to predict future effects, the hazard assessment is a preliminary activity that
helps to define assessment endpoints. This is accomplished by determining which environmental
components might be potentially exposed to the stressor and how they may be affected. The
hazard assessment is also used to determine whether a hazard exists by comparing the
magnitudes of expected environmental concentrations to toxicological test endpoints for a
contaminant. The hazard assessment is the most commonly used methodology for analyzing the
effects of chemicals on the natural environment.
Assessments to Support Resource Management
Managers offish, game, forest, and land resources conduct ecological assessments to support
their decisions. These assessments typically analyze the consequences of human activities (i.e.,
harvesting) on the health and sustainability of these natural resources. Managers are typically
concerned with species-specific and site- or region-specific issues. There are often large bodies
of data available. As a result, this type of ecological assessment tends to rely heavily on
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statistical and mathematical models. In addition, many resource managers are developing
probabilistic model predictions to support their management decisions.
Environmental Evaluations of New Chemicals, Biologicals, and Pesticides
These evaluations provide information regarding whether a proposed substance or biological
organism will adversely affect non-target organisms.
Scientific Issue Assessment
Scientific issue assessments analyze the potential future ecological effects arising from
environmental concerns that have not yet been proven to have an effect on the environment (e.g.,
global warming).
Assessment of Environmental Change
These types of assessments are performed because of observed environmental changes. They
attempt to explain the nature and extent of the effects by determining the probable or possible
causes. For example, it has been observed that there is a hole in the stratospheric ozone layer
over Antarctica, reducing the protection provided to the earth's surface from damaging
ultraviolet radiation. Scientists are attempting to determine the size and timing of the hole (i.e.,
does the size fluctuate, remain constant, or consistently increase/decrease). They are also
examining possible causes (e.g., atmospheric pollutants or natural cyclic phenomena).
CERCLA Risk Assessment
This type of assessment is conducted at Comprehensive Environmental Response, Compensation
and Liability Act (CERCLA) hazardous waste sites. These assessments determine the
probability that adverse ecological effects are occurring due to a release of hazardous wastes
from the site.
Natural Resource Damage Assessment
Natural Resource Damage Assessments (NRDAs) take both an ecological and economic
standpoint, because standard methodologies have been promulgated by the Department of the
Interior (DOI) and the National Oceanic and Atmospheric Administration (NOAA) to both assess
the injury to an ecological resource and evaluate the economic damages. In 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
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assessment is similar to an ecological risk assessment in that the researcher must document the
source of the stressor and the exposure pathways.
Section 4 goes into more detail on Ecological Risk Assessments. The purpose of the section is to
help the economist better understand the scientific framework for analysis and type of
information that may be generated through an ecological assessment.
References and Further Reading
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.
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.
U.S. EPA. 1998. Guidelines for Ecological Risk Assessment, EPA Document 630/R-95/002B.
Washington, DC.
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. 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-CO-0070. January.
U.S. EPA. 1992a. Framework for Ecological Risk Assessment Washington, DC: U.S.
Environmental Protection Agency, Risk Assessment Forum. EPA/630/R-92/001. February.
U.S. EPA, Office of Policy Planning and Evaluation. 1992b. Biological Populations as
Indicators of Environmental Change, EPA Document 230-R-92-011. Washington, DC.
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4,0 THE ECOLOGICAL RISK ASSESSMENT PROCESS
An ecological risk assessment often provides the basis for the economic analysis of the benefits
associated with an action that impacts the environment. As such, it is important to understand
the context from which the results of the ecological risk assessment are generated. With this
understanding, the analyst can look for opportunities to leverage the data collection and better
apply the results, therefore improving the economic benefit assessment. This section provides a
basic understanding of the methodology of ecological risk assessment by highlighting the major
steps and issues. The reader should refer to the appropriate references for a thorough
understanding.
4.1 OVERVIEW OF 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 (EPA, 1998). Some examples of
ecological stressors are:
Q Physical
• Erosion
• Heat
• Turbidity
• Impoundments
• Habitat alterations
Chemical
• Hazardous substances (e.g., pesticides,
industrial wastes)
• Salinity
• Air pollutants (CO, NOX, ozone)
Biological
Risk assessments may be conducted to
determine the risks associated with:
A stressor — what are the risks to the
ecosystem associated with the air
emissions from an incinerator?
An effect — what are the risks to an
ecosystem resulting from a decline in
rainbow trout populations?
An ecological component — what are
the risks for one fish species versus
another from increased siltation?
• Disease-causing organisms (Pfiesteria, diatoms)
• Genetically-engineered microorganisms
• Non-native species (kudzu, zebra mussels)
Ecological effects assessments include magnitude, duration, spatial distribution, time to
recovery, and other relevant parameters. Ecological risk assessments may be predictive (i.e.,
they estimate whether changes in the ecological components are likely to occur as a result of an
anticipated event or stressor), or they may be retrospective, (i.e., what is the probability that a
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Stressors in Waquoit Bay, MA
Multiple chemical, biological and physical
stressors are believed to be affecting Waquoit
Bay. Principal stressors include:
Chemical
Physical
Nutrients, such as phosphorus and
nitrogen
Toxic chemical contaminants
Suspended and resuspended
sediments with increased water
turbidity
Physical alteration of habitat
Altered water flow in a watershed
Biological
Finfish harvest pressure
Eelgrass wasting disease caused by
slime mold
past event caused this present problem. A
predictive risk assessment may take the
form of modeling the effects of a known
or anticipated stressor, such as
atmospheric nitrogen levels in the
Chesapeake Bay. The assessment depends
on previously collected data from similar
events and applying it to a new situation.
In a retrospective ecological risk
assessment, an effect may be well defined,
such as decreased bird populations, but
determining the cause of the decline is the
goal of the risk assessment. In this
example, it may be a pesticide that is
poisoning the bird, or it may be loss of
nesting habitat. The risk assessment is
estimating the probability that any one
stressor or combination of stressors is
causing the decline in the bird population..
It is important to note that adverse effects
can be expressed quantitatively (e.g., as a
probability, such as one in a million)
and/or qualitatively (e.g., low, medium, or
high). In addition, the uncertainty
associated with the probability is typically provided. Ecological risk assessments may be
conducted in a tiered or incremental fashion beginning with an assessment of a single stressor
on a single receptor and progressing on to an evaluation of numerous effects caused by several
stressors at many levels.
Steps to an Ecological Risk Assessment
An ecological risk assessment starts with planning. During the first stage, assessors and
managers determine the need and scope of an ecological risk assessment. It is during this stage
that societal and political issues are factored into the formulation of the problem. All key
participants should be involved in planning the risk assessment. This 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 consensus on the goals, scope, and timing of the
ecological risk assessment. If the goal is to achieve a regulatory mandate, proper planning will
ensure that the outcomes meet the objective and support implementing management decisions.
Scientific data gathering and analysis required for the ecological risk assessment occurs once:
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Q Objectives of the risk assessment have been defined (including criteria for
success);
Q Goals for ecological values have been established;
Q A range of options have been developed;
Q Focus and scope of the assessment have been agreed upon; and
Q Resources to conduct the assessment have been provided (EPA, 1998).
EPA defines three phases for conducting ecological risk assessments (EPA 1998):
Q Phase 1 — Problem Formulation;
Q Phase 2 — Analysis (exposure assessment and ecological effects
characterization); and
Q Phase 3 — Risk Characterization.
The overall framework for the ecological risk assessment is presented in Exhibit 8. The
scientific conduct required for each of these phases is discussed in the following subsections.
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Exhibit 8
ECOLOGICAL RISK ASSESSMENT FRAMEWORK
Planning
(Risk Assessor/
Risk Manager
Dialogue)
PROBLEM FORMATION
Assessment
Endppints
ANALYSIS
Characterization
of
Exposure
Characterization
of
Ecological
Effects
/_
RISK
CHARACTERIZATION
Risk
Estimation
Risk
Description
Communicating Results
to the Risk Manager
I
Risk Management
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4.2 PHASE I: PROBLEM FORMULATION
Problem formulation is a formal process for generating and evaluating preliminary hypotheses
about why ecological effects have occurred, or may occur, from human activities (EPA 1998).
It provides a foundation upon which the entire ecological risk assessment depends. Problem
formulation involves the development of three products: assessment endpoints, a conceptual
model, and an analysis plan. Because problem formulation is inherently interactive and
iterative, substantial re-evaluation is expected to occur throughout the risk assessment process.
Assessment Endpoints
An ecological risk assessment must begin with
well-defined assessment endpoints. An
assessment endpoint describes both the valued
ecological entity (e.g., a species, ecological
resource, habitat type, or community) and the
function or measure of particular concern (e.g.,
reproductive success, production per unit area,
surface area coverage, or biodiversity) (EPA,
1998). Assessment endpoints are defined by
characteristics of the exposed environment are
important in the decision-making process.
Factors can include ecological and societal
relevance, susceptibility to known or potential
stressors, ecological or economical value, and
ability to be measured.
When supporting an ecological benefit
analysis, risk assessors should collaborate with ^^^~^~^~"~
economists and policy-makers to help identify
the links between the ecological assessment endpoints and the economic benefit endpoints (see
section on Interdisciplinary Coordination). For example, when assessing the impact of an
insecticide known to kill red-winged blackbirds, the assessment might also examine the
secondary impacts on bald-eagles, because of their high societal value. Bald-eagles are
impacted, because they eat the poisoned blackbirds. By identifying the link to a highly visible
and highly valued ecological resource, such as the bald eagle, the economist will be better able
to assign a monetary value to the impact of the lesser known (often lower trophic level) species
or a change in the function of the ecosystem, that is otherwise very difficult to value.
Example Assessment Endpoint
A decline in salmon population would be an
appropriate choice as an assessment
endpoint for a study measuring the impact
of the construction of a hydroelectric dam
on a river in the Pacific Northwest. Salmon
have ecological relevance, because they are
a food source for many aquatic and
terrestrial species, and they eat many
aquatic invertebrates. Salmon are also
sensitive to changes in sedimentation, water
temperature, and substrate pebble size.
Most importantly, salmon are valued by
society as a source of food, a form of
recreational fishing, and for their
ceremonial and symbolic significance to
Native Americans .
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Assessment Endpoints in Waquoit Bay Risk Assessment
During the problem formulation phase, eight assessment endpoints were selected to
represent the estuarine and freshwater components of the Bay:
• Eelgrass abundance and distribution;
• Resident and nursery estuarine finfish diversity and abundance;
• Eelgrass-dependent estuarine benthic invertebrate community diversity and
abundance;
• Trout and alewife migratory fish reproduction;
• Riverine benthic invertebrate community structure and function;
• Freshwater pond trophic status;
• Water-dependent wildlife feeding and nesting habitat; and
• Bacterial and contaminant content offish and shellfish.
Assessment endpoints may be defined for both structural and functional aspects of an ecological
resource. For example, biological endpoints may be any level of organization ranging from a
single individual to an entire ecosystem. Most endpoints, however, are 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.
Ecological assessment endpoints are not
confined to effects on structural aspects of
the ecosystem. The endpoint might also be a
change in a functional process, such as a
decrease in the primary productivity of a
forest, a decrease in the oxygen content of a
pond, or the loss of nitrification capacity in
contaminated soil. Loss of specific habitat,
such as a decrease in wetland acreage
resulting from the construction of a dam,
might be a combined structural and
functional endpoint for some ecological risk
assessments (Bartell et al., 1992).
Assessment Endpoint Selection
Assessment endpoints should:
Be ecologically relevant (i.e., do not use
southern pine as an endpoint for
northern forests),
Be susceptible to the stressor (i.e., if
examining effects of sedimentation, do
not use catfish as an indicator organism;
use rainbow trout), and
Represent societal values (i.e., assess
impacts on an endangered bird species
rather than a perceived nuisance
species).
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Given the complexity of ecological systems and components, there is no universal list from
which to select appropriate assessment endpoints. There are, however, three criteria that an
assessment endpoint should satisfy (EPA 1998): ecological relevance, susceptibility to the
stressor, and societal relevance.
Stakeholder Involvement in Establishing Management Goals
Waquoit Bay provides an excellent example of how stakeholder involvement can be
instrumental in developing management goals for an ecological risk assessment. It has
been generally agreed by all involved parties that the bay is changing — eelgrass is
disappearing and is being replace 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.
Ecological Relevance
The ecological relevance of an endpoint refers to its importance in relation to other components
of a specific community or ecosystem. This includes a decrease in the population of a certain
insect species that affects the populations or productivity of another species. For example, the
honeybee is ecologically significant in some pasture lands, because it is a pollinator of many
plants. Effects on certain species offish may be selected as an assessment endpoint because of
the critical role the fish species plays in maintaining the functional integrity of the ecosystem
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Example Assessment Endpoint
The change in the population of lake
trout represents an appropriate
assessment endpoint that links a receptor
(lake trout in the Great Lakes) to an
effect (competition from zebra mussels)
resulting from exposure to a stressor (a
change in the concentration of
polychlorinated biphenyls). The
endpoint, change in the lake trout
population, is clearly defined, easily
measured, and reflects the societal and
ecological impacts of changes in
pollution levels. It may be quite
difficult, however, to identify
appropriate assessments eridpoints to
evaluate the impacts on some important
ecological qualities and functions, such
as biodiversity or ecological integrity.
(e.g., top consumer in the aquatic food chain).
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 rely on the algae to oxygenate the water.
Increased deer populations resulting from lack of
predators may cause excessive damage or death to
young trees upon which the deer feed.
Susceptibility to the Ecological Stressor
The assessment endpoint must be based on a link
establishing that the ecological component (or
receptor) being examined is susceptible to and may
be exposed to an ecological stressor.
Susceptibility is based on the vulnerability of the
ecological component to the stressor and is
influenced by many factors including the mode of
action of the stressor, the life history of the
receptor (for biological receptors), and the life
stage of the organism. For example, survival of an
endangered species may be at greater risk from the
loss of a single individual than would a healthy,
numerous species, because the healthy species population could rapidly recover. 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.
Susceptibility is based on exposure. Some ecological components may be located close to the
stressor, may be exposed for a longer time, or at a greater intensity. The results of exposure may
not be immediately evident. Susceptibility may be obvious some time after exposure to the
stressor has ceased, as is the case with many birth defects that skip a generation. Susceptibility
may be greater for a secondary effect of the stressor. For example, application of a pesticide
may kill the target organisms such as mosquitos in a stagnant pond (primary effect) but 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). Multiple stressors
may also increase the susceptibility of the ecological component. A seal with a virus may be
weakened and thus become easier prey for a shark.. Because ecosystems contain many variables,
both living and non-living, determining all vulnerable receptors and the extent of their exposure
may be a difficult task.
Relevance to Management Goals (including Economic Relevance)
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 decision-makers such as
clean air in parklands. In some cases, such an evaluation may be difficult if the ecological
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and/or economic significance of the
component is not well understood. It may
even be disliked by the public. For example,
blue-green algae in tidal marshes often have
noxious odors but they form the basis for a
food web that ultimately supports a sport fish
population. An appropriate assessment
endpoint for a risk assessment to determine
the effects of damming an inlet to the marsh
would be to determine the effects on the sport
fish population, not the blue-green algae.
Ecological risk assessments are most effective
when they measure the impacts to those
ecological components that are used directly
by humans, such as sport fish, groundwater,
or timber land. Assessment endpoints might
also measure the impacts to those ecological
components and processes that indirectly
benefit humans, such as water filtration,
climate control, or flood protection. What is
actually measured may be different from what
is important to the economic valuation study,
but the relationship between the two elements should be clearly defined. In many cases, the
selection of an assessment endpoint is pre-determined 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 helpful if changes to
the assessment endpoint can be predicted or measured, particularly if a dose-response
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. In many cases, a dose-response
relationship may be impossible to quantify although the 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.
Establishing Susceptibility to the Stressor
In Waquoit Bay, submerged aquatic
vegetation (e.g., eelgrass) is a sensitive
indicator of eutrophication, particularly
nitrogen loading. Increased nitrogen
concentrations result in increased algal
growth. Excessive algal growth can result
in direct shading of the plants, preventing
photosynthesis. As the eelgrass dies off, the
amount of preferred habitat available for
juvenile scallops is reduced. The juvenile
scallops themselves, however, are not
directly affected by the nitrogen
concentration in the water. Therefore, an
appropriate assessment endpoint for
nitrogen loading would be the extent of
eelgrass in the Bay rather than the number
of scallops.
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Linking the Assessment Endpoints to the Management Goal in Waquoit Bay
The goal 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 its presence is an indicator of good water
quality. The disappearance of eelgrass may be because of 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 measured.
Scallop abundance in the Bay is not a good assessment endpoint, because the scallop
population is naturally highly variable, therefore, changes in population abundance would
not necessarily reflect the effect of a stressor.
Conceptual Models
Risk assessors explore potential interactions between stressors and the assessment endpoints by
developing a conceptual model (EPA, 1998). The conceptual model links the stressor, exposure
pathways, ecological receptors, and ecological 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).
Conceptual models include two principal
components: risk hypotheses and a
conceptual model diagram.
Risk hypotheses are statements of
assumptions based on available information
(EPA, 1998). They describe predicted
relationships between stressors, exposure,
and assessment endpoint response. Early
conceptual models are intended to be broad
in scope, identifying as many potential
relationships as possible. As more
Example of a Risk Hypothesis
Bald eagles had been found dead or dying.
Analysis of their stomach contents indicates
that they had consumed smaller birds. The
risk hypothesis may be that the eagles were
eating poisoned birds who were easy prey
and that the poison was a pesticide that had
recently been applied to local fields. Such
was the case for the pesticide carbofuran.
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information is incorporated, the plausibility of specific hypotheses is determined. The most
appropriate risk hypotheses are identified for subsequent evaluation in the analysis phase of the
risk assessment. Risk hypotheses do not necessarily involve statistical testing of null and
alternative hypotheses or any particular analytical approach. They are intended to put the risk
assessment problems in perspective by indicating what is known about the risk and what
relationships need to be evaluated. Risk hypothesis may be developed for specific effects,
predictions of the effects of a stressor (e.g., what are the range of effects from a chemical), an
ecosystem, or ecosystem component, such as a watershed or a deer population.
A conceptual model diagram (see Exhibit 9) is a useful way to visually express the
relationships described by the risk hypotheses. Conceptual model diagrams can communicate
important exposure pathways in a clear and concise way. Risk assessors can use these
diagrams, along with the risk hypotheses, to select the pathways that will be evaluated in the
analysis phase of the ecological risk assessment. These diagrams and hypotheses also are useful
tools to aid communication with economists and policy makers. 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.
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Analysis Plan
In an analysis plan, risk assessors describe the data and measures that will be used to evaluate
the risk hypotheses, i.e., the effects on the assessment endpoints (EPA, 1996). Measures are
identified for: exposure, ecosystem and receptor characteristics, and effects. Exposure to
stressors may be quantified or estimated (e.g., how much enters the environment and how it is
distributed, including its possible degradation or modification). Measures of ecosystem and
receptor characteristics identify important life history traits that affect the receptors' potential
exposure or the response of assessment endpoints to the stressors (e.g., reproductive cycles,
migration patterns, and habitat types). Measures of effects quantify the response of the
receptors to 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: what is the endpoint
(eelgrass) and its attributes (extent of habitat). It is the latter aspect that must be measured
(quantitatively or qualitatively and directly or indirectly). Eelgrass habitat may be measured by
aerial photography, thus effects on the habitat are quantitatively determined. However, for
some assessment endpoints such as songbird populations (entity) and their decline (attribute) as
a result of pesticide ingestion, it may be difficult to count how many birds are actually affected
by the pesticide if they are able to fly away before dying. In many cases, surrogate measures of
effects must be used (e.g., toxic effects on other birds in a laboratory setting).
Assessment Endpoints and Measures
An ecological risk assessment is to be conducted for a pulp mill on a Pacific northwest
river. One assessment endpoint may be Coho salmon breeding success and fry survival.
Possible measures of the effects of the mill 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 populations, reproductive cycles, and laboratory tests for
reproduction, growth, and mortality. Measures of exposure may include: contaminant
concentrations in water, sediment, and fish, and dissolved oxygen levels in the water.
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Exhibit 9
WAQUOIT BAY ESTUARY PARTIAL CONCEPTUAL MODEL
Fertilizer
Application
Septic
Systems
Sewage
Treatment
Plants
Surface Water
and Sediment
Nutrients
1
tation
I
1
Shad in
|
Reduced Scallop Population
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4.3 PHASE II: ANALYSIS PHASE
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 (EPA, 1996). 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 conceptual model and analysis plan provide the basis for the analysis
phase. Based on the conceptual model, the risk assessors should know which stressors and
ecological effects are the focus of the investigation. In the analysis plan, the risk assessors
identified the information needed to perform the analysis phase. The analysis plan and
conceptual model were conducted in Phase I.
The analysis phase is composed of two activities: characterization of exposure and
characterization of ecological effects (EPA, 1998). These assessments are usually conducted
simultaneously and interaction between the scientists conducting them is recommended.
Exposure Assessment
Characterization of exposure through an exposure assessment identifies the source of the
stressor, the distribution of the stressor in the environment, and the contact or co-occurrence of
the stressor with ecological receptors. The exposure assessment should identify the source of
the stressor and the complete pathway by which it is acting upon the 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. Exposure analysis may
start with the source when it is known, but in cases where the source is unknown, the analysis
may attempt to link the contact of the stressor with the receptor (e.g., chemical residues in fish
tissues) to a source.
Describe the Source of the Stressor
As part of describing the source of the stressor, the risk assessors identify where the stressor
originates. The source of the stressor can be the place were the stressor is released into the
environment (e.g., a smoke stack, a farmer's field) or the action that produces the stressor (e.g.,
dredging). In some assessments, the original source no longer exists, and the source is defined
as the current origin of the stressor. For example, the source of polychlorinated biphenyls
(PCBs) may be defined as contaminated sediments, because the industrial plant that produced
the contaminants no longer operates and the contaminants have become embedded in the
sediments.
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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.
In addition to establishing the original or current source of the stressor, the stressor itself should
be described in terms of its temporal and spatial scale. Several factors that may be considered in
describing a stressor include:
Intensity - How much of the stressor is in the environment and at what levels 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
(decreasing atmospheric ozone)?
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)?
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 distance (stratospheric ozone).
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Stressor generation may be presented
quite simply. For example:
Hazardous chemicals (stressors) from
an outfall (source) are released for 10
minutes every 24 hours for 250 days
of the year with releases, totaling
200,000 tons of hazardous chemicals
per year. The temperature of the
outfall is 35 °C. This example
illustrates hazardous chemicals and
temperature as two potential
stressors.
Many stressors have natural
counterparts 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.
Nitrogen Loading in the Chesapeake Bay
The Chesapeake Bay is eutrophic, causing excess
algal growth and declines in fish populations.
Several possible sources of excess nutrients have
been indicated:
• Atmospheric deposition
• Run-off from agricultural land
• Industrial waste streams
Although fertilizer runoff is the most obvious source
of the pollution, atmospheric deposition, which may
originate many miles from the watershed, has been
demonstrated to be a significant loading factor. All
or some of these sources may be considered in the
ecological risk assessment depending on the
management goals.
Describe the Distribution of the Stressor in the Environment
The spatial and temporal distribution of chemical stressor(s) in the environment are described
by evaluating the pathways they 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 portions of the environment, the assessors
describe the temporal and spatial distribution of the changed environment (e.g., 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 may have the ability to
reproduce in suitable environments, and do not need a medium to move from one area to
another. Therefore, when identifying the exposure pathways for biological stressors, both active
and passive modes of distribution need to be considered.
The environmental fate of a Stressor depends on several aspects:
Distribution: Once in the environment, where does the stressor go? Stressors may be
released or formed from various environmental media. A pollutant released to water may
partition to the sediment or remain in the water column. Different physical forms of a
stressor may partition to different media.
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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 methyl mercury, 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 have been
directly exposed.
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
can be found in the environment. The physical and chemical characteristics of both the
stressor and the receiving environment determine its transport and distribution.
Degradation or Transformation: Degradation may occur via biotic processes
(metabolism or Bioaccumulation), 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).
Identifying the distribution, transport, degrada-
tion, or transformation processes to which a
stressor is subject provides an indication of
how much 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.
The formation and subsequent distribution of
secondary stressors may be important depend-
ing on the objectives of the assessment. For
chemicals, the evaluation of secondary
stressors usually focuses on metabolites or
degradation products. Disturbance of the
environment can also lead to secondary
stressors. Several methods may be used to
understand the distribution and environmental fate of a stressor and characterize the potential
exposure of specific receptors to the stressor. Ideally, direct monitoring by collecting and
Examples of Secondary Stressors
Chemical: Aldicarb is toxic to mammals
but not very persistent in the environment.
However, it is rapidly degraded to aldicarb
sulfone, which is toxic, very persistent, and
moves through the soil to the groundwater
where it may remain for years.
Physical: Dredging of a waterway not only
causes loss of habitat for the organisms at
the site of the activity, but may result in
severe turbidity of the water.
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analyzing environmental (including biological) samples is preferred. Monitoring should be
conducted so that appropriate spatial and temporal samples are taken. Monitoring will help
define the area over which the stressor may be acting and any changes in the level of stressor
(including its degradation products) over time.
Where monitoring information is lacking or difficult to obtain, models may be used to estimate
the 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 and chemical properties, as well as the environment of concern to
characterize the amount and extent of stressor that may be acting on a receptor. Typically, a
combination of monitoring and modeling is used to determine the stressor levels.
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 would be
possible. However, if a pathway from source to receptor cannot at least be hypothesized, then it
may be assumed that the receptor will not be affected by the stressor. For example, if a toxic
chemical released to surface waters from an industrial outfall is converted quickly to non-toxic
compounds, fish downstream may not be exposed to the chemical and the pathway is
incomplete. If the chemical conversion is dependent on the acidity of the water, however, and
the acidity of the outfall is altered so that the chemical is no longer quickly degraded, the
downstream fish may be exposed and the pathway is complete.
Pathways may also be direct (the links between source of the stressor and receptor are easy to
establish) or indirect (the stressor may act upon one receptor that in turn causes effects in a
second stressor). The example of the toxic chemical used above is a direct pathway — the
stressor causes adverse effects in the receptor. However, in Waquoit Bay there are several
indirect pathways. For example, the nitrogen loading to the bay does not immediately affect the
scallop population. Rather the effects of the nitrogen loading on scallops is seen as a secondary
effect of phytoplankton growth, decline in eelgrass habitat, and finally, a decrease in scallops.
Consequently, a comprehensive exposure assessment must include as much information as
possible about the source of the stressor, its fate and transport in the environment, receiving
media, and availability of the receptor(s), both primary and secondary.
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
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consideration for some ecological stressors is the level of biological organization that is
affected, and whether the stressor is acting directly or indirectly upon the receptor. For
example, a stressor may cause adverse effects at multiple feeding levels within a community
(i.e., multiple trophic levels), and these effects may increase or decrease at higher trophic levels,
depending on the nature of the stressor. A classic example of an ecological stressor causing
effects at higher trophic levels of the food chain is the bioaccumulation of DDT in birds and
mammals. On the other hand, a biological stressor such as the tobacco mosaic virus, may have
a profound impact on vegetation in a community, but no immediate impact on insects that are
feeding on the affected plants unless their food source is eliminated. In general, an ecological
risk assessment will attempt to capture the effects of a stressor at all trophic levels by the
assessment of food-web interactions. Other examples of adverse effects on ecological
components that may be measured in an assessment are sickness, death, decreases in
reproduction rates and productivity in populations, decreases in community biodiversity, and
changes in predator-prey relationships.
It is also important to know the characteristics of the potential receptors in the exposed area.
For example:
Q Are they present on a permanent basis (e.g., trees), or are they migratory (e.g.,
many species of birds)?
Q Can and do receptors avoid exposure (i.e., are they capable of movement (e.g.,
some biota may be able to move from contaminated areas but trees, soil, and
water bodies can not))?
Q What are population parameters, such as the size and distribution of the receptors?
Q Is population is in a growth or decline phase?
Q Are stressor effects likely to occur when the population is particularly vulnerable
(e.g., during molting or when nesting)?
Q What are physical and temporal parameters, such as seasonal and diurnal changes
in non-biological receptors (e.g., does the lake freeze in the winter?)?
Exposure can be described in several different ways, depending on how the stressor causes
adverse effects:
Co-occurrence of the stressor with receptors. Co-occurrence is particularly useful for
evaluating stressors that can cause effects without actually contacting ecological
receptors. For example, whooping cranes prefer sandbars with unobstructed views in
rivers for their nesting areas. Thus, manmade obstructions, such as bridges, can interfere
with the nesting behavior of whooping cranes without actually contacting the birds. Co-
occurrence is evaluated by comparing the distribution of the stressor with the distribution
of the ecological receptor. For example, by overlaying two maps, one showing the
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placement of bridges, the other areas historically used by nesting whooping cranes, the
assessor can evaluate the impact of manmade obstructions on crane nesting patterns.
Contact of a stressor with receptors. Most 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. Contact is a function of the amount of a stressor in an
environmental medium and the activities or behaviors that bring receptors into contact
with the stressor. For chemicals, contact is quantified as the amount ingested, inhaled, or
applied to the skin. Chemicals are also absorbed over the gills offish and aquatic
invertebrates.
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. Uptake is usually assessed by modifying
the estimate of contact to account for how much of the stressor is available.
Establishing a baseline for the natural state of the individual, population, community, or
ecosystem is vital for determining the extent of effects caused by the stressor and ensuring that
appropriate endpoints have been identified and characterized.
When the analyses and supporting documentation have been completed, the exposure
assessment should provide a description of the amount of 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.
Characterizing Exposure Effects
An ecological effects characterization describes the relationship between the stressor and the
magnitude of the resulting ecological effects. The ecological effects characterization indicates
what effect the stressor may have on various receptors and the levels that may elicit different
responses (i.e., the stressor-response relationship). Many stressors do not affect all receptors in
the same way or at the same levels. In Waquoit Bay, for example, nitrogen loading is a
significant stressor. Increased nitrogen levels in the Bay result in excessive phytoplankton
growth that has two effects: (1) decreased eelgrass habitat because the phytoplankton shade the
eelgrass preventing photosynthesis, and (2) decreased oxygen levels in the water that cause
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. The
ecological effects characterization involves three steps: determining the stressor-response
relationship(s), evaluating causality, and linking the measure of effects to assessment endpoints.
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Determine the Stressor-Response Relationship
Evaluating ecological risks requires an understanding of the relationship between stressor levels
and the resulting ecological effects. Stressor-response analysis (often called dose-response
analysis for chemical stressors) is a method typically used. Stressor- response analysis is often
used for chemical stressors such as toxic substances. However, the technique may 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 success. This type
of analysis is particularly valuable, because it measures the different effects that a stressor may
have at many levels. For example, a slight increase in temperature (stressor) in a given stream
may lead to a significant decline in the trout population (response), but only a minor one for
algae population. If the temperature continues to increase, however, the algae will also
eventually experience a decline in population. Therefore, it is important that responses be
evaluated. The level of exposure may help determine the critical stressor-response
relationship(s).
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. For
some chemicals, the effects may be reversible once exposure ceases.
Stressor-response analysis often provides a quantitative characterization of the stressor and
effect. Examples of quantitative characterization for chemical stressors include acute toxicity
values, such as lethal dose or concentration (LD50 or LC50). The values represent the dose or
concentration that will kill half of the test organisms in a specified time). Other toxicity values
include maximum acceptable toxicant concentration (MATC), or inhibition of photosynthesis.
Stressor-response relationships are not always linear (e.g., an increase in 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, algae may not cause a decrease in
productivity until the turbidity of the water column stops the light from penetrating the surface
and photosynthesis cannot occur. Some stressors may have disproportionate effects on
receptors if the receptors are already subject to another stressor. If deer are starving because of
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deep snow hiding their food, the introduction of wolves may reduce the deer population by
greater numbers than expected.
Stressor-response information is typically obtained from laboratory or field studies.
For some stressors, a quantitative characterization may.be difficult to develop. In these cases, a
qualitative characterization may be used. For example, the assessor could note that a marsh area
was receiving fewer visits from migratory fowl, because of the lack of food source, without
attempting to enumerate the number of birds during a particular time period.
Evaluate Causality
Some ecological risk assessments begin
with a known stressor whose ecological
effects are well understood. Risk assess-
ments, however, are driven by observed
adverse ecological effects, such as bird or
fish kills. In these instances, risk assessors
may identify possible stressors responsible
for the ecological effects. Then they
attempt to determine which stressor is the
actual cause. 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
been subjected to epidemics of a
distemper- like 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 (EPA, 1992b).
Therefore, while the assessment endpoint
can be identified for the receptors (i.e., a
change in the seal population), the
stressors can only be suggested and the
possible impacts described without a
quantitative analysis. Thus, although it may be possible to put an economic value on the change
in the seal population, the change in the value cannot be definitively linked to a specific stressor.
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
Blackbumian warbler, eat insects and live in
the interior forest where they breed. They
migrate south in the winter, following their
food supply.
The risk hypothesis is that possible causes of
the population decline are forest fragmentation
in North America and tropical deforestation in
South America. Forest fragmentation results
in larger forest boundary areas that are not
conducive to breeding for these birds, which
are subsequently more vulnerable to predators
(including other birds) and parasites.
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.
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The following criteria may be used for evaluating causality (EPA, 1998):
G Criteria strongly affirming causality:
• Strength of association
• Predictive performance
• Demonstration of a stressor-response relationship
• Consistency of association
Q Criteria providing a basis for rejecting causality:
Inconsistency in association
Temporal incompatibility
Factual implausibility
Other relevant criteria:
• Specificity of association
• Theoretical and biological plausibility
Establishing Causality for Declines in Fish Populations
Declines in fish populations have been reported for many species in waters around the world.
Although some declines can be traced to specific causes, in many cases, no particular
stressor has been implicated as the sole cause. Among the possible causes offish population
declines are:
» Pathogens such as the bacteria Pfiesteria piscicida
• Declines in food sources (as a result of natural cycles or from manmade
disturbances)
• Changes in water chemistry (natural or as a result of pollution)
• Loss of habitat for shelter or breeding
• Physical obstructions (e.g., salmon cannot return upriver because of dams)
• Over-fishing
• Competition by more successful species (e.g., decline in indigenous fish species
in the Great Lakes as a result of the introduction of zebra mussels)
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.
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The following are examples of extrapolations that risk assessors may use to link measures of
effect to assessment endpoints (EPA, 1998):
Q Between similar organisms (e.g., bluegill to rainbow trout);
Q Between responses (e.g., mortality to growth or reproduction);
Q Between different sources of data (e.g., laboratory to field data);
Q Between geographic areas (e.g., northeastern U.S. to northwestern U.S.);
Q Between spatial scales (e.g., stream to river); and
Q 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), risk
assessors identify the extrapolations required between assessment endpoints and measures of
effect. Decisions about specific extrapolations are usually based on the scope and nature of the
risk 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 ecological impacts from
the stressor(s) known to be present in the
ecosystem. Based on this information the
risk assessors can then determine if the
stressor(s) warrants attention and, if so,
what can be done to prevent further effects.
Characterizing Uncertainty
Uncertainty evaluation is an ongoing issue
throughout the analysis phase. The purpose
of an uncertainty analysis is to formally
Uncertainty Factors
Uncertainty factors may be quantitative or
qualitative depending on their application.
In the development of a conceptual model
for the risk 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
defined (e.g., use of coyotes rather than
wolves). A pathway may not be clearly
defined from the source of the stressor to the
stressor. For example, a species of bird may
have an impaired reproduction rate. The
risk assessment assumes that loss of habitat
from timber cutting is the stressor, when it
may be that the birds are exhibiting
reproductive effects as a result of runoff
from the timber cutting exposing
contaminated soil. The pathway for the
ecological risk assessment therefore is not
certain.
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recognize the incomplete knowledge that the ecological risk assessment is constructed upon and
to explain these 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 connected with an ecological risk assessment should investigate (at a
minimum) the uncertainty or potential error associated with any extrapolations that are made
during the ecological risk assessment. The uncertainty 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. Among the extrapolations that may be needed are:
Q Extrapolation from one exposure duration to another (e.g., short-term (acute) to
long-term(chronic)) (see Uncertainty Factors box);
Q Extrapolation from one species to another (e.g, using alewives as a surrogate for
rainbow trout);
Q Extrapolation from one trophic level to another (e.g., using midges as a surrogate
for praying mantis);
Q Extrapolation from one response to another (e.g., egg shell thinning as a
surrogate for decreased reproductive success);
Q Extrapolation from one stressor to another (e.g., use of structure-activity
relationships to estimate the toxicity of one chemical based on the toxicity of a
structurally similar chemical);
Q Extrapolation from laboratory to field (e.g., using effects on laboratory animals
or artificial ecosystem as a surrogate for natural populations or whole
ecosystems); and
Q Extrapolation from one ecosystem to another (e.g., effect on freshwater lake as a
surrogate for an estuarine bay).
In addition to the uncertainty in extrapolation there are other causes of uncertainty in an
ecological risk assessment. These include, but are not limited to:
Q Variability in samples of data;
Q Inability to obtain appropriate samples (this may be of concern if the organism is
endangered or difficult to identify or collect);
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Q Lack of knowledge about multiple chemical effects, interactive synergistic
effects, and counteractive synergistic effects; and
Q Non-linear behavior of complex systems.
Because quantitative measures of uncertainty are often difficult (and sometimes impossible) to
provide, 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 strengths and weaknesses, (see also
Interdisciplinary Coordination section.)
Methods for analyzing and describing the uncertainty associated with the ecological risk
assessment range from simple to complex. When presenting analysis results as a point estimate,
classical statistical methods, such as confidence limits and percentiles, can effectively describe
the uncertainty associated with these estimates. When a modeling approach is used, sensitivity
analyses and Monte Carlo analyses can be used to evaluate how sensitive the model's outputs
are to variations in the input values (see also Issues section). Uncertainty propagation can also
be analyzed to examine how the uncertainty associated with individual parameters affects the
overall uncertainty associated with the conclusions of the risk assessment.
4.4 PHASE III: RISK CHARACTERIZATION
Risk characterization is the final phase of ecological risk assessment. The goals of this phase
are to (EPA, 1998):
• Use the results of the analysis phase to estimate the risk of ecological effects
represented by the endpoint identified in the problem formulation phase;
• Interpret the risk estimate (including an assessment of the uncertainty associated
with the estimate); and
• Report the results.
Risk Estimation
Once the ecological effects of a stressor have been characterized (through the ecological effects
characterization performed in Phase II), that information is combined with the exposure
assessment to provide an indication of the ecological risks and associated uncertainty. In other
words, the risk assessors determine the likelihood of adverse effects resulting from the presence
of the defined stressor. This determination may be qualitative or quantitative. Where specific
data are lacking, such as in the example of declining seal populations, the assessors may need to
exercise professional judgment to determine the most or least likely risks.
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Several approaches are available for quantitatively estimating ecological risks: compare single-
point estimates, incorporate the entire stressor-response relationship, and decide based on the
process models.
Compare Single-Point Estimates
When sufficient data is available to develop
quantitative exposure and effects estimates,
the simplest approach for comparing the
estimates is to use a ratio of the two
numbers. Typically, the ratio (or quotient) is
expressed as an exposure concentration
divided by an effects concentration.
Quotients are commonly used for chemical
stressors, where reference or benchmark
toxicity values are widely available as
measures of effect. In most cases, the
quotient method does not explicitly consider
uncertainty (e.g., does not address the
uncertainty associated with the extrapolation
from the tested species to the species or
community of concern). The uncertainty
associated with the single-point estimates
can, however, be addressed by providing a
statement of the likelihood that the exposure
point estimate exceeds the effects point
estimate.
Hazard Quotients
If the 96-hour LC50 of a chemical in the
fathead minnow is 1 mg/L and the
concentration of the chemical in the river is
0.001 mg/L, then the ratio of the exposure
concentration to the effect concentration
(the hazard quotient) is 0.001 or 1/1000.
The ratio suggests that the chemical
concentration in the river is of relatively low
risk to fathead minnows. A stressor with a
hazard quotient of less than one is expected
to pose little risk to the receptor (assuming
the tested organism is at least as sensitive as
any indigenous organism). Whereas, a
hazard quotient of greater than one is
expected to pose a risk to exposed
organisms.
This approach is used for many regulatory risk assessments. For example, if it has been
determined that the salinity in a river to which wastes are discharged exceeds the relevant
ambient water quality criteria, then it is assumed that the biological organisms in that river are at
risk for adverse effects. For single, large short-term (acute) exceedences, this may result in
death. However for lower, longer exceedences (chronic) this may mean stunted growth or other
non-lethal effects. The regulatory action associated with this determination is that the discharge
of wastes should be reduced so that the level of the stressor is below the ambient water quality
criteria.
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Estimates Incorporating the Entire Stressor-Response Relationship
Stressor-Response Relationships
Gypsy moths have been shown to decimate
many forests in the northeastern United
States. A small number of gypsy moth
larvae may cause minor damage to the
foliage on some trees. However, a larger
infestation may result in stunted tree growth
or even tree death if the larvae eat enough
leaves where trees cannot sustain their
photosynthetic requirements. The level of
gypsy moth activity may be directly related
to tree damage, up to and including death.
In cases where sufficient data are available to
indicate a range or distribution of effects,
such as a stressor-response curve (where
increasing levels of the stressor produce more
severe adverse effects), the risk assessment
may statistically analyze the effects and
compare them to a range of exposures.
Monte Carlo simulations or other approaches
for incorporating uncertainty may be used to
indicate a mean and standard deviation of the
data. This approach also provides a spectrum
of effects and exposures against which
management decisions may be based. For
example, when developing ambient water
quality standards it is helpful to determine
worst-case or statistical confidence limits to
be predictive of the most sensitive members of a population. The greater the variability in the
exposure or stressor-response relationship, the greater the number of risk estimates. This
variability may provide a more realistic approach to risk assessment, as both high-end and low-
end exposures may be considered. Risk estimates can also be made for average or healthy
populations (e.g., adults) in addition to sensitive populations (e.g., young animals).
Estimates Based on Process Models
Process models are mathematical expressions that represent our understanding of the
mechanistic operation of a system under evaluation. They can be useful tools both in the
analysis phase and the risk characterization phase of the ecological risk assessment. A major
advantage of using process models for risk estimation 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 species-level effects to population and ecosystem levels. These models may also be of use
in estimating indirect effects on the assessment endpoints and the probable rate of recovery.
A variety of these 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.
The methods described above are those most commonly used for quantitatively estimating
ecological risks. Slight variations might be employed, such as comparing point estimates of
effects with cumulative exposure distributions. Other less commonly used techniques, that are
described in the EPA guidelines, involve incorporating variability in the exposure or effects
estimates to describe the risks to highly exposed or extremely sensitive receptors. Field
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observational studies also can serve as a risk estimation technique, although causal relationships
between the stressors and the effects must first be established.
While quantitative models are often useful in providing ecologists and economists with numeric
estimates of potential effects, it may only be possible to provide a qualitative description. For
example, clear cutting a forest may be expressed as the number of trees cut or acres cleared (one
assessment endpoint), but it may not be possible to determine the effects in bird populations as a
result of this action.
Risk Description
After the ecological risks have been estimated, risk assessors need to integrate and interpret the
available information into conclusions about the risks to the assessment endpoints. Risk
descriptions provide technical narrative supporting the risk estimates as well as a framework for
interpreting the estimates. Professional judgment is required to assess the various endpoints and
identify those which are likely to experience the greatest short-term and long-term effects. In
some cases, adjustments may be required if the data indicates that the proposed assessment
endpoints are of less concern than others identified during the assessment process.
The reliability of the risk assessment must also be discussed. For example, modeling may
indicate that one effect is of greater magnitude than another although empirical evidence may
suggest otherwise. This problem arises because of the imprecision of the quantity and validity
of the input data or parameters of the model. Models may not incorporate all of the influential
parameters. Sometimes, other factors affecting the ecosystem may not be immediately evident.
The risk assessors must evaluate all effects and determine which options deserve further
consideration. It may be necessary to conduct additional monitoring or testing to validate the
model and confirm the existing data.
Reporting Risks
A risk assessment report may be briefer extensive depending on the nature of the assessment.
It is important that the information be presented clearly and concisely. The outputs of the
ecological risk assessment may become the inputs for future management decisions as well as
the economic valuation phase of the benefit assessment. The findings of the risk assessment
report should be discussed with economists, policy makers, and other users of the risk
assessment.
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Hanemann, W.M. 1991. "Willingness to Pay and Willingness to Accept: How Much Can They
Differ?" American Economic Review. 81(3): 635-647.
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 Land. New York, New York: Columbia
University Press.
Willig, R. 1976. "Consumer Surplus Without Apology." American Economic Review 66(4):
589-597.
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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 Background Theory section for a
discussion on defining the
economic value of ecological
resources). This section
discusses the various types of
goods and services and offers
their taxonomy, which may be
useful in developing a compre-
hensive list of specific economic
benefit endpoints for the
ecological benefit analysis. The
proposed taxonomy for
generally characterizing the
goods and services provided by
ecological resources is presented
in Exhibit 12. Direct Use Indirect Use
Some of the goods and services
provided by ecological
resources are obvious because
they are directly used or enjoyed
by society, such as the fish
Exhibit 12
Proposed Taxonomy of Goods and Services
Provided by Ecological Resources
Good or Service
Non-Use Value
Use Value
Direct, Ivferket Direct, Non-Market Indirect, Non-Nferket
Good or Service GoodorService GoodorService
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provided by a fishery, the timber/lumber provided by a forest, or the 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 "uses" of 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 because it is
more difficult to obtain information on the "price" of the service and the number of people
enjoying the service (i.e., benefitting from the resource through a specific use), because the
goods or services are not sold through markets.
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. 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, valued
through that relationship (see Interdisciplinary Coordination section). These types of services,
which are not bought and sold through markets, are referred to as indirect, non-market uses.
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. Another example of a non-use value would be bequest value,
which refers to the value people place on a maintaining a resource for future generations.
The value of a change to a specific ecological resource can be estimated, in part, by measuring
the change in the value of the direct, market uses and direct, non-market uses 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 and, therefore, 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.
For example, a partial list of the goods and services society derives from birds might include:
• Food source (direct, market use);
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• Hunting, bird watching, and contributing to the aesthetic environment for hikers,
campers, anglers, and other recreationists (direct, non-market use);
• Component to an ecosystem that supports or provides other goods and services
and contribute to maintaining biodiversity (indirect, non-market use); and
• 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 describe the various methods available for estimating the
economic value of changes to these goods and services to society.
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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
(i.e., change in welfare) 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 value of the change experienced by the ecological resource.
Examples of Direct, Market Uses Provided by Ecological Resources
G Food Source
• Fish (specific species) - commercial fishery
• Crops (specific type: corn, beans, apples, etc.) - commercial and home
production
• Animal (fowl, deer, etc.) - commercial consumption
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Q Building Materials
• Timber (specific species)
• Stone
Q Fuel
• Timber (specific species)
Coal
Oil
Q Drinking Water Supply
• Ground water reservoir
• Surface water reservoir
Q Medicine
Q Chemicals/Minerals
Valuing Direct Market Uses
There are a number of market-based approaches that may be useful in estimating the value or
change in value of a direct market use provided by an ecological resource. In most cases, a
market-based 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 events affecting the resource that do not affect the population dynamics or
the overall level of use of the resource (i.e., 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. 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. These
second-best approaches do not directly reflect welfare changes as described in background.
Specific revealed preference techniques that value marketrbased goods include:
• Market-Price and Supply/Demand Relationships
• Market-Based Valuation Approaches
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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
G Fishing
• Recreational Fishing (specific species, area)
• Subsistence Fishing (specific species, area)
Q Beach Use (sunbathing, swimming, walking)
Q Recreational Hunting (specific species) - for sport and/or personal consumption
Q Bird Watching (general, specific species)
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Q Tourism
Q Boating
Q Hiking/Camping
Q Animal Viewing, Photography, Feeding (general, specific species)
Q Sightseeing
Q Aesthetic Value
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 offish 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 direct, non-market uses include:
Revealed Preference Methods:
• Hedonic Price Methodologies
• Travel Cost Methodologies
• Random Utility Models
Stated Preference Methods:
• Contingent Valuation
• Contingent Activity and Combining Contingent Valuation with Other Approaches
• Conjoint Analysis and Contingent Ranking
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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
Q Flood Control
Q Storm Water Treatment
Q Ground Water Recharge
Q Climate Control
Q Pollution Mitigation
Q Wave Buffering
Q Soil Generation
Q Nutrient Cycling
Q Habitat Value
Q 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
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resource. These methods assume that people demonstrate, or reveal, the value they place on the
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 indicate the 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 the indirect, non-market uses
include:
Revealed Preference Methods:
• Hedonic Price Methodologies
• Replacement/Alternative Cost
• Avoidance Expenditures
Stated Preference Methods:
• Contingent Valuation
• Contingent Activity and Combining Contingent Valuation with Other
Approaches
• Conjoint Analysis and Contingent Ranking
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6.2.4 NON-MARKET, NON-USE VALUES
Non-market non-use values of an ecological resource include the value 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
Q Scarcity Value
Q Option Value
Q Existence Value
Q Cultural/Historical Value
Q Intrinsic Value
Q Bequest Value
Q 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 non-
use 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
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people's responses to proposed choices or questions regarding the value they place on certain
ecological resources (known as contingent valuation). Determining the total non-market non-use
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.
Only one technique is applicable for estimating non-market, non-use values:
• Contingent Valuation
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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 in the
goods and services provided. 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 technique (more detailed
descriptions are provided in later sections).
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 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
willingness-to-pay) or must monetary values be inferred based on a model of
individual behavior (i.e., indirect estimation of willingness-to-pay)?
Exhibit 13 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.
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Exhibit 13
Categorization of Valuation Techniques
Direct Estimation of WTP
Indirect Estimation of WTP
Revealed Preferences
Approach
Market Price/Quantity
(Estimated Supply/Demand)
Market Simulation Models
User Fees
Replacement Costs
Travel Cost Studies
Random Utility Model
Hedonic Studies
Avoidance Expenditures
Referendum Voting
Stated Preferences
Approach
Contingent Valuation Studies
Contingent Ranking
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
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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.
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
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• 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
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:
• Contingent Ranking: 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; or
• Contingent Activity: 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 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.
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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.
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 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.
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6.3.1 MARKET-PRICE AND SUPPLY/DEMAND RELATIONSHIPS
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 offish), 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. For these cases, 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, (see the Background Theory section for additional
discussion on using estimated supply and demand relationships to estimate the benefits of an
action.) 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).
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Market Simulation Models
More recently, 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.
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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.
>• It may be difficult to correctly estimate demand and/or supply relationships if
limited data on prices and quantities is available.
>• It may be difficult to separate the supply and demand effects and to isolate the
effects of the environmental change.
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.
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6.3.2 MARKET-BASED VALUATION APPROACHES
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 they value 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 offish breeding and stocking programs to estimate the value offish
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
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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 project
has been implemented or if society has demonstrated their 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.
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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.
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 14 presents some of the possible measures for estimating the benefits of reducing
pollutant levels using defensive expenditures:
Exhibit 14
Estimating the Benefits of Reducing Pollutants Using Defensive Expenditures
Pollutant
Air Pollution
Water Pollution
Hazardous Waste
Noise Pollution
Radon in well water
Radon in Soil Underneath House
Defensive Expenditure Measures
Clean or repaint exterior of house; install air
purifiers or new air conditioners; visit the
doctor more frequently; move away from
pollution source
New well; bottled water; water purifiers;
move away
Similar to both water and air pollution
depending upon medium by which hazardous
waste affects households
Storm windows; thicker walls; move away
Filter or aerate water; bottled water; increase
house air ventilation; move away
Ventilate crawlspace of house; seal
foundation of house; use thicker concrete in
basement; increase house air ventilation;
move away
Source: Bartik (1988).
Referendum
Referendums 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
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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. 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; and
• Different jurisdictions vote on different levels of a good.
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. 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.
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
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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).
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 public goods - 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 l/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.
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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.
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6.3.3 TRAVEL COST METHODOLOGIES
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 (or the average demand price) can be used as an estimate of the average
willingness-to-pay for a trip (Loomis, 1993).
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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.1
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 underestimate 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 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 (Bishop and Heberlein, 1979).
>• 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.
>• Environmental quality and other site characteristics may be difficult to describe in
quantitative terms.
'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).
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Data Requirements
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 a/., 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. 1991. "Recreation." in Braden, J.B. and C.D.
Kolstad,eds. 1991. Measuring the Demand for Environmental Quality. North-Holland,
Amsterdam: Elsevier Science Publishers.
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.
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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.
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.
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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 and well ordered, 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
(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.
U S EPA Headquarters Library
Mai! code 3201
1200 Pennsylvania Avenue NW
Washington DC 20460
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>• 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.
>• Specification as with all techniques and estimation procedures can have a
significant impact on benefit estimates.
>• 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.
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 sociodemographic information). Additionally, accurate measurement of the
characteristics of the alternative sites is necessary.
Example Random Utility Study (Englin et at., 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 was 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 marginal 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. ^^^^^^^^
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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.
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6.3.5 HEDONIC PRICE AND HEDONIC WAGE
METHODOLOGIES
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.
Figure 1
Graphic Illustration of a Hedonic Price Equation
for an Environmental Quality Attribute (Pearce and Turner, 1990)
Property A
Price, PF
I Slope of PP
Pollution Level Environmental Quality, E
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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-to-
pay, 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«/., 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 Vz 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.
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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. Opportunity cost, defined as the difference between the marginal
implicit price of distance from a salt pond for houses with average attribute levels in the salt
pond region versus the northern section of the study region, was estimated to be $509.
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).
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
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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).
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 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
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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.
>• 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.
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. This 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, TJ. 1988. "Measuring the Benefits of Amenity Improvements in Hedonic Price
Motels." 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.
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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., P.M. Fritz, and T. Vukina. 1997. "Hog Operations, Environmental Effects, and
Residential Property Values." Land Economics 73(1): 114-124.
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. HarperCollins
Publisher.
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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). 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.
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 sociodemographic (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).
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Generally, these criteria include (at a minimum):
• Interview a large sample of the affected population;
• Achieve a high response rate;
• 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.
Advantages
X 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.
X Comparisons of benefit measures from 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.
X Contingent valuation methods are the only methods available for estimating non-use
values (e.g., existence values).
X The willingness-to-pay estimates from contingent valuation include both the use and
nonuse value of the good or service.
X Survey-based contingent valuation methods can capture respondents' attitudinal and
behavioral information that are not available in other non-survey based valuation
techniques.
Disadvantages
X 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
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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. 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 in a contingent valuation study 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.
>• 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 (Bishop and
Heberlein, 1979).
Example Contingent Valuation Study (Whittington etal., 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
Houston-Galveston 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.
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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.
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.
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
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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 respondent is then 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 double-
bounded, approach increases the statistical efficiency of the valuation estimate and reduces the
necessary sample size.
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Advantages
X 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.
Example Contingent Valuation Study (Carson et a/., 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 willingness-to-pay
per individual was estimate to range from $50.61 to $52.81.
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Iterative Bidding Games
Generally bidding games are conducted through personal interviews where the interviewer
iteratively questions the respondent. 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-to-accept)
a program or plan that will affect their well-being, as well as sociodemographic 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
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contingent valuation survey, the quality of the results depends in 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
Arrow, K., R. Solow, P.R. Portney, E.E. Learner, R. Radner, and H. Schuman. 1993. "Report of
the NOAA Panel on Contingent Valuation." Federal Register, January, Vol. 58(10): 4601-4614.
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.
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.
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.
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.
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6.3.7 COMBINING CONTINGENT VALUATION WITH OTHER
APPROACHES: CONTINGENT ACTIVITY
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.
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Example of Combining Contingent Valuation and Travel Cost Data (Cameron, 1992)
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. The in-person survey, conducted by the Texas
Department of Parks and Wildlife, asked 3,366 respondents:
(1) 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;
(2) How much they will spend on their current fishing trip; and
(3) 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
restrictions on days of access. Specifically, Cameron estimated that a 10 percent reduction in
fishing days would result in a $35 loss in welfare, on average. 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.
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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 etal, 1995).
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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 logit 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
X Conjoint analysis allows the valuation of an action as a whole and the various
attributes or effects of the action.
X 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).
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>• 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 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.
Disadvantages
>• Respondents may find some trade-offs difficult to evaluate or unfamiliar to them.
X 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 *r a/., 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.
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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. Environmental Protection Agency. Report 230-05-83-001. Washington
D.C.
Freeman, A.M., IIL 1993. The Measurement of Environmental and Resource Values: Theory
and Methods. Washington, D.C.: Resources for the Future.
Johnson, F. R., W.R 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.
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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.
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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 (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 (Kirehhoff era/., 1997).
There are three commonly used benefit transfer techniques:
• Mean unit value transfer
• Adjusted unit value transfer
• 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
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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
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 (EPA, 1995):
Q Differences in attributes of the policy case or site, level of use, or in the
socioeconomic characteristics of users affected by the change;
Q Differences in the environmental policy, change, or resulting effects; and
Q 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
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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
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.
X 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
X 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.
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Example Benefits Transfer Study (Bowen et aL, 1993; 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.
This document provides information to allow researchers to begin exploring the viability of
benefits transfer for their particular study. Specifically, this document includes a valuation
database, which presents a selection of resource values reported in the literature, as well as a
bibliographic database, to help the reader locate specific articles and reports of interest.
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.
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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. LaJolla,
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
Plan (CCMP) Planning and Implementation, A National Estuary Program Environmental
Valuation Handbook. Washington, D.C.
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7.0 ISSUES AFFECTING THE ECONOMIC VALUATION OF
ECOLOGICAL BENEFITS
This section identifies and briefly discusses some additional issues that may affect the economic
benefit analysis, most of which are not discussed in earlier sections of this document. These
issues include:
• Uncertainty and Variability;
Aggregation;
• Discounting; and
- • Equity.
Discounting, and other issues including baselines, uncertainty, non-monetized effects, equity, and
valuing lethal risks, are currently being examined by EPA's Office of Policy, Planing, and
Evaluation (OPPE) in support of the preparation of a revised guidance for preparing Economic
Impact Analyses and Regulatory Impact Analyses. These guidelines, therefore, should be
evaluated in the course of a comprehensive ecological benefits analysis.
7.1 UNCERTAINTY AND VARIABILITY
The variability and uncertainty associated with specific estimates is an important consideration in
a thorough benefits assessment. Variability and uncertainty are introduced in many ways
including estimating emissions changes, modeling the fate and transport of emissions (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.
Distributional information from both the risk assessment and the economic valuation study
should be carried through in the final benefits assessment.
There are several treatments of variability and uncertainty available that can be applied in the
estimation of benefits. The appropriate approach to characterizing and quantifying the degree of
variability and uncertainty associated with a specific estimate will depend on the objectives of the
analysis and the needs of the decision-makers. Depending on the particular valuation approach
used to develop a benefits estimate, the uncertainty and variability associated with the results of
that approach might be addressed by:
• Presenting the benefits estimates as ranges based on a plausible set of input values
(e.g., risks, values);
• Conducting sensitivity analyses to examine the potential variation in the benefits
estimates under different assumptions regarding the level of effects;
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• Using Monte Carlo analyses or other probabilistic techniques to estimate a
probability distribution for the output (e.g., benefits);
• Discussing and/or incorporating expert judgement regarding the potential range of
effects and/or benefits (e.g., Delphi methods); and/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 unqualified. Thus, qualitative descriptions of the limitations and known
omissions, biases, and data gaps are also an important component of a thorough benefits analysis.
7.2 AGGREGATION
Numerous issues arise in aggregating individual willingness-to-pay (WTP) estimates to develop a
social or national-level benefits estimate. Although many of these issues require additional
theoretical deliberation to determine the most appropriate approach, below are three aggregation
issues that warrant consideration in the applied context of estimating benefits as described in this
document:
• How to sum benefits across benefit endpoints;
• How to address potential double-counting when using multiple techniques to
measure the WTP for changes in related benefit endpoints or overlapping effects;
and
• How to determine the affected population for calculating social WTP and how to
sum WTP over the affected population.
7.3 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 the time period for comparison can have significant
effects on the magnitude of the benefits estimate and the conclusions of the benefits assessment,
especially if the 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
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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 intergenerational 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, annualized the costs and benefits or accumulate benefits (and costs)
forward to some future time period.
7.4 DISTRIBUTIONAL AND EQUITY ANALYSES
Distributional and equity analyses examine the distribution of changes incurred 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. Thus, under certain economic structures, 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 a moral perspective.
Therefore, in conducting an ecological benefit analysis it is important to pay attention to the
distribution of costs and benefits (i.e., track who in society is benefitting from the change and
who is not). 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.
Decision-makers then use the results of the distributional or equity analysis, along with the
results of the ecological benefit analysis, socio-economic impact analysis, cost analysis, other
analyses and moral, legal, and/or philosophical considerations, to evaluate the proposed policy or
action.
References and Further Reading
U.S. EPA. 1997. Discounting in Environmental Policy Evaluation, Draft Final Report.
Prepared by Frank Arnold, Fran Sussman, and Leland Deck for the 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 Office of Policy, Planning, and Evaluation.
May 23, 1997.
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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 Office of Policy,
Planning, and Evaluation. May 27, 1997.
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8.0 REFERENCES AND DATA BASES
8.1 ECOLOGICAL REFERENCES AND FURTHER READING
Bartell, S.M., R.H. Gardner, and R.V. O'Neill. 1992. Ecological Risk Estimation. Boca Raton,
FL: 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.
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).
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.
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.
U.S. EPA. 1998. Guidelines for Ecological Risk Assessment. EPA Document 63 0/R-95/002B.
Washington, DC.
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.
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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-CO-0070. January.
U.S. EPA. 1992a. Framework for Ecological Risk Assessment. Washington, DC: U.S.
Environmental Protection Agency, Risk Assessment Forum. EPA/630/R-92/001. February.
U.S. EPA, Office of Policy Planning and Evaluation. 1992b. Biological Populations as
Indicators of Environmental Change. EPA Document 230-R-92-011. Washington, DC.
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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.
Arrow, K., R. Solow, P.R. Portney, E.E. Learner, R. Radner, and H. Schuman. 1993. "Report of
the NOAA Panel on Contingent Valuation." Federal Register, January 15, 58(10): 4601-4614.
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 a/., 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, DJ. and J.R. Kahn, eds. 1996. The Contingent Valuation of Environmental
Resources: Methodological Issues and Research Needs. Brookfield, Vermont: Edgar Elgar
Publishing Ltd.
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.
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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.
Daily, G., ed. 1997. Nature's Services: Societal Dependence on Natural Ecosystems.
Washington, D.C.: Island Press.
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.
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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.
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.
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.
Haniey, N. and C.L. Spash. 1993. Cost Benefit Analysis and the Environment. Brookfield,
Vermont: Edward Elgar Publishing Limited.
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.
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.
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Loomis, J.B. 1993. Integrated Public Lands Management: Principles and Applications to
National Forests, Parks, Wildlife Refuges, and BLMLand. 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.
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.
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.
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 MJ. 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.
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Palmquist, R.B., P.M. Fritz, and T. Vukina. 1997. "Hog Operations, Environmental Effects, and
Residential Property Values." Land Economics 73(1): 114-124.
Pearce, D.W. and R.K. Turner. 1990. Economics of Natural Resources and the Environment.
Maryland: The Johns Hopkins University Press.
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.
Rowe, R.W. 1985. Valuing Marine Recreational Fishing on the Pacific Coast. LaJolla,
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. HarperCollins
Publisher.
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. Discounting in Environmental Policy Evaluation, Draft Final Report,
Prepared by Frank Arnold, Fran Sussman, and Leland Deck for the 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 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 Office of Policy,
Planning, and Evaluation. May 27, 1997.
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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, 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.
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.
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8.3 ILLUSTRATIVE SAMPLE OF DATA SOURCES
This section provides an illustrative sample of data sources that might be used to support an
economic assessment of ecological benefits. Potential types of data sources include compilations
of results from primary analyses, reference publications, data bases, and programs/projects
involved with the collection of potential relevant information.
Program or Project
American Fisheries Society Special Publication
24: 1992 Investigation and Valuation of Fish Kills
(and Sourcebook)
Benefits Transfer of Outdoor Recreation Demand
Studies, 1968-1988.
A Bibliography of Contingent Valuation Studies
and Papers.
Biological Status and Trends Program
Current Fishery Statistics
Economic and Environmental Principles and
Guidelines for Water and Related Land Resources
Implementation Studies, 1983.
Economics Valuation of Wetlands. Discussion
Paper 065
The Environmental Economics Database
(a database of existing natural resource or
environmental resource value estimates)
Environmental Valuation Reference Inventory
(EVRI)
EMAP Agricultural Lands Resource Group
Forest Inventory Analysis
Contact Organization
American Fisheries Society
Bethesda, Maryland
Article by Walsh, R.G., D.M. Johnson, and J.R. McKean.
1992 in Water Resources Research 28(3)
Carson, R.T., N. Carson, A. Alberini, N. Flores, and J.
Wright. 1993. Natural Resource Damage Assessment, Inc.,
La Jolla, California.
United States Geological Survey/BRD
300 National Center
12201 Sunrise Valley Drive
Reston, VA20192
National Marine Fisheries Service, U.S. Department of
Commerce, Washington D.C.
U.S. Water Resources Council
Washington D.C.
Prepared for American Petroleum Institute by Anderson,
R., and M. Rockel. 1991.
U.S. Environmental Protection Agency, Office of Policy,
Economic Analysis and Research Branch
(has not been updated since 1994)
U.S. Environmental Protection Agency
Environment Canada
U.S. Department of Agriculture
http://www.epa.gov/emfjulte/html/datal/agroland/index/ht
ml
U.S. Department of Agriculture, Forest Service
Northeastern Experiment Station
5 Radnor Corporate Center, Suite 200
Radnor, PA
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Program or Project
Forest Ecosystem Health Project
Gap Analysis Program (GAP)
Master Environmental Library (MEL)
National Environmental Research Parks (NERP)
National Environmental Monitoring and Research
Initiative
Natural Resources Inventory
TEMS Database (Terrestrial Ecosystems
Monitoring Sites)
Contact Organization
U.S. Department of Agriculture, Forest Service
U.S. Environmental Protection Agency
U.S. Geological Survey/BRD
National GAP Office
530 Ashbury St. Suite 1
Moscow, ID 83843
U.S. Department of Defense/Defense Modeling and
Simulation Office
U.S. Department of Energy
U.S. Environmental Protection Agency
U.S. Department of Agriculture/National Resources
Conservation Service (NRCS)
United Nations, Global Environmental Monitoring System
Program Activity Center
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8.4 DATA BASE OF ECOLOGICAL RESOURCE VALUES
This section provides a sample of the information that will be available through the electronic
document using a key-word searchable data base. This information is provided to guide the
reader to a sample of potentially useful studies and is not intended to provide the analyst with
values to be used directly in an economic valuation study. Readers should obtain and thoroughly
review any study cited to evaluate the quality of the estimate and the appropriateness of the value
to their specific analysis.
Example of Valuation Database Entries for keyword "forest":
Location
Measurement Unit
Study Value
i,
Basis for }
Estimate i
Walsh, R.G., R.D. Bjonback, R.A. Aiken, and D.H. Rosenthal. 1990. Estimating the Public |
Benefits of Protecting Forest Quality. Journal of Environmental Management 30:175-189. i
Colorado
Willingness-to-pay per resident
household per year for forest quality
protection programs
$14.00
Contingent I
Valuation j
Dwyer, J.F., G.L. Peterson and A.J. Darragh. 1983. Estimating the Value of Urban Forests
Using the Travel Cost Method. Journal of Arboriculture 9(7): 182-185.
Chicago
WTP by Chicago metropolitan
residents
$8.68
Travel Cost i
Hagen, D., J. Vincent, and P. Welle. 1992. Benefits of Preserving Old-Growth Forests and the !
Spotted Owl. Contemporary Policy Issues 1 0(April): 1 3-25. i
Oregon
Household WTP for owl protection
$36.91
Contingent
Valuation
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8.5 BIBLIOGRAPHIC DATA BASE
This section provides a sample of the information that will be available through the electronic
document using a key-word searchable data base.
Example of Bibliographic Entries for keyword "forest":
1. Adger, N. and M. Whitby. Accounting for the Impact of Agriculture and Forestry on
Environmental Quality European Economic Review 35(2-3):629-641, 1991
2. CAAA Retrospective Study. "Comparison of Morbidity, Visibility, and Forest Valuation
Studies to Contingent Valuation Guidelines" Draft, September 30, 1993
3. CAAA Retrospective Study. Approach to Environmental Benefits Assessment to Support
Clean Air Act Section 812 Analysis, November 6, 1992
4. Crocker, T.D. On the Value of the Condition of a Forest Stock. Land Economics
61(3):244-254, 1985
5. Dwyer, J.F., G.L. Peterson, and A.J. Darragh. Estimating the Value of Urban Forests Using
the Travel Cost Method. Journal of Arboriculture 9(7): 182-185, 1983
6. Dwyer, J.F., H.W. Schroeder, J.J. Louviere, and D.H. Anderson. Urbanities Willingness-
to-pay for Trees and Forests in Recreation Areas. Journal of Arboriculture
15(10):247-252, 1989
7. Energy and Resource Consultants, Inc. "The Benefits of Air Pollution Control in
California" California Air Resources Board Chapter 8: Forest Benefits, 1986
8. Englin, J.E. and R. Mendelsohn. A Hedonic Travel Cost Analysis for Valuation of
Multiple Components of Site Quality: The Recreation Value of Forest Management.
Journal of Environmental Economics and Management 21:275-290, 1991
9. Haynes, R.W. Economic Evaluation of Acidic Deposition on Forests. A paper presented at
the El-Economic, Policy and Law-Working Group Session at the SAF National
Convention held at Rochester, NY on October 16-19, 1988
10. Lockwood, M., J. Loomis, and T. DeLacy. A Contingent Valuation Survey and Benefit
Cost Analysis of Forest Preservation in East Gippsland, Australia. Journal of
Environmental Management 38(3):233, 1993
U.S. EPA Headquarters Library
Mail code 3201
1200 Pennsylvania Avenue NW
Washington DC 20460
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11. Loomis, John B. Integrated Public Lands Management: Principles and Applications to
National Forests, Parks, Wildlife Refuges, and ELM Land. Columbia University Press,
New York, New York, 1993
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