In order to evaluate how an action affects human well-being, quantification of changes, negative or positive, in ecosystem services is necessary. Our experience has shown that quantified estimates of ecosystem services sufficient for environmental decision-making can be obtained through the use of the valuation approaches developed and refined under the United States natural resource damage assessment (NRDA) process. Within NRDA, these approaches are used to balance compensatory restoration with adverse impacts so as to maintain the flow of natural resource services provided to the public. The approaches include economic-based methods to quantify the gains or losses in ecological and human use services associated with an action that affects the environment. It is within this context that the use of these economic-based methods is considered and takes advantage of the experience gained within the NRDA process over the past 20 years. We propose the use of these ecological and human use quantification methods to evaluate changes in ecological and human use services associated with green practices. In this manner, green practices can be compared regarding their ability to provide a net ecosystem service benefit and hence, their potential sustainability. The proposed approach does not necessitate that a restoration-based offset be incorporated into an alternative practice in order to demonstrate sustainability. The need and appropriateness of an offset should be discussed among the stakeholders involved.
In general, the term “natural resource services” as used within the NRDA process, is equivalent to the term “ecosystem services”. Natural resource services are classified as either: (1) ecological services or (2) human use services [
22,
23,
24]. Human use services are further grouped into direct (e.g., consumptive) and indirect (e.g., passive or non-consumptive) uses (
Figure 7). It should be recognized that natural resources and associated habitats serve as the foundation from which human use services flow. The classification of ecosystem services used within the NRDA process and presented in
Figure 7 helps to minimize the potential for double counting of ecosystem service flows that is apparent with the Millennium Assessment (MA) classification of ecosystem services [
16]. In the MA classification [
16], ecosystem services are segregated into supporting services from which regulating, provisioning, and cultural services flow (
Figure 1). If service flows are quantified and valued according to this classification, there is a potential for considerable double counting of service values. For example, supporting services are those services, such as primary production, that support the production of provisioning, regulating, and cultural services. When both supportive and provisioning services, for example, are accounted for in the same analysis, the supporting services are counted twice, both directly and then indirectly through the production of provisioning services [
25]. Double counting can also occur across provisioning, regulating, and cultural services as all categories are valued directly on their own and indirectly through their production and support of final ecosystem goods and services [
25]. Unless ecological production functions for services are known and well specified, it is difficult to appropriately categorize, quantify, and value all service flows. Therefore, due to double-counting issues associated with quantifying ecosystem services within the MA classification, ecosystem service flows are quantified within the NESA approach according to the service classifications presented in
Figure 7.
Figure 7.
Classification of ecosystem services within the natural resource damage assessment (NRDA) process [
24].
Within the United States, the need for compensatory restoration has resulted in the growth of the service-to-service approach to develop appropriately scaled compensation for contaminant releases [
26,
27,
28,
29,
30]. The concept of the service-to-service approach is that compensation for lost ecosystem services can be conducted through the provision of services into the future. This concept inherently incorporates tradeoffs to make the public “whole”. The tradeoff being the ability to provide equivalent habitat and/or human use services through restoration (
i.e., offset project) to compensate for those services lost. The provision of ecological services can be conducted through restoration activities such as the acquisition, creation, enhancement or preservation of habitat. If required and determined to be appropriate, it is preferred that the restoration (
i.e., the provision of habitat and human use services) be “in-kind”, that is, replace the same type and quality of habitat impacted. In addition, it is preferred that the restoration be spatially connected (e.g., in the same watershed) to the impacted habitat. However, in several cases, compensatory restoration has been “out of kind” and not spatially linked to the watershed impact [
30,
31]. In these cases, the regulatory agencies preferred one habitat type over another or made a linkage between the restored resource and the injured resource (e.g., if a release impacted migratory waterfowl, habitat restoration that supports that species along its flyway has been accepted). These decisions were the basis of tradeoffs made on behalf of the public by the resource agencies involved. In regards to human use services, compensatory restoration has included projects that create human use value such as the creation of boat ramps, parks,
etc.It should be recognized that the incorporation of offsets is not always appropriate in the demonstration of environmental sustainability since ecosystem services are spatially explicit and it may not be possible to replace ecosystem services in some areas. In some cases, given the geographical location, certain ecosystem services may have a higher priority when compared to others, based upon their overall contribution to human well-being in that particular area. That is, the importance and priority associated with a specific ecosystem service can be different based upon the specific location to which that benefit is provided.
The practical approach presented herein incorporates the ecological and human use quantification methods commonly used in the service-to-service approaches as a means to quantify changes in ecosystem service flows given an action that affects the environment. In addition, the impact and benefit metrics used within the methods consider both non-monetary and monetary ecosystem service flow metrics. These metrics can be incorporated into the evaluation of ecosystem service flows associated with green practices in order to establish their overall sustainability compared to a baseline condition. In some cases, the financial benefits may drive the development of a particular project and offsets may need to be developed if appropriate. As such, consistent with the definition of sustainability, offsets have the potential to be incorporated into the alternatives analysis of green practices and the quantification of net ecosystem service flows for the purposes of demonstrating sustainability. As stated above, offsets should be considered on a case-by-case basis.
Figure 8.
Demonstrating Potential Environmental Sustainability using a Net Ecosystem Service Analysis Approach.
3.1.1.1. Ecological Service Quantification
As development continues, the public, through government policies, commonly makes trade-offs associated with actions affecting the environment. A good example is related to the 404 wetland permitting process in the United States. Under this process, project impacts to ecological services (e.g., wetlands) and associated human use services (e.g., water filtration) can be compensated for through the provision of wetland habitat to offset the impact (e.g., development impacts to wetland habitat). The policy provides a basis for making environmental tradeoffs. That is, impacts are allowed to occur provided the public is willing to accept the proposed compensation. This is consistent with dynamic efficiency and incorporates the notion that the benefits associated with the compensatory wetland restoration, acquisition, and/or enhancement are appropriately scaled to the level of habitat service loss so that the benefits are greater than or equal to the wetland losses. A method for comparing the ecological habitat service flows is provided in King and Adler [
33]. This concept led to the development of the habitat equivalency analysis (HEA) method for determining appropriately scaled compensatory restoration [
21,
24,
26,
27,
28,
29,
30].
Because many ecological habitat services are not traded in the marketplace, they do not have a direct monetary value. The HEA approach is a service-to-service equivalency approach that evaluates ecological habitat service losses and gains based on non-monetary metrics over time. It is important to recognize that ecosystem services are not static measurements but represent a flow of benefits over time as represented in
Figure 3,
Figure 4,
Figure 5,
Figure 6. Primary applications of HEA have been associated with balancing the negative ecological habitat impacts of an action with the positive ecological habitat benefits associated with a compensatory restoration action [
29]. The balancing (
i.e., equivalency) concept is based on the assumption that the public can be compensated for past and/or projected estimated losses in ecological habitat services through the provision of similar services of the same type in the future assuming the value the public places on losses and gains in ecological habitat services is the same [
30].
This also assumes that indirect (passive and non-consumptive) values are incorporated within the ecological service value as well. While restrictive, an implication of this assumption is that it is not necessary to attempt to place a monetary value on the injured and restored ecological service flows, That is, despite the fact that HEA is an economic model, it uses ecological metrics to measure the flow of services. This provides an advantage over other methods of determining equivalent value because it can reduce the data and analysis requirements. The HEA methodology is flexible and can be adapted to individual sites and situations.
The HEA approach uses an environmental metric to measure changes to ecological habitat services and focuses on quantifying the area (e.g., hectares, acres) and level of impact over time in units typically represented as service-hectare-years (SHY’s) or service-acre-years (SAY’s). However, other metrics can be used and developed appropriate to the site and the key services being evaluated. HEA allows for service losses associated with an impact and service gains associated with restoration to be quantified in the same units so that offsetting mitigation, if needed, can be scaled equivalent to the adverse impact. The formal HEA methodology concurrently calculates both the level of impact and, given the characteristics of potential restoration, the appropriate level of restoration to compensate for those losses. However, the HEA methodology can be adapted to independently quantify either the relative impacts or benefits, independent from each other, associated with actions. This is the typical application within a NESA. For example, in comparing between various green practices, HEA might only be used to estimate relative project benefits. There may not be impacts to quantify. In other cases, a green practice may create potential service losses and gains concurrently. In others, the HEA may demonstrate an ecosystem service loss associated with an action. In order to illustrate the use of HEA to estimate ecological habitat service flows, one can apply, given the characteristics of either the impact or benefit of an action, the net change in ecological habitat service flows over time. By example, the HEA methodology could be used to calculate the areas of the various curves (shaded areas) presented in
Figure 4,
Figure 5,
Figure 6 in units of SAY’s, SHYs, or other alternative site specific metrics (e.g., services per linear distance of stream). An overview of the information, technical approach, and input parameters required to conduct habitat equivalency analysis is detailed in published articles and government sponsored reports [
21,
26,
27,
28,
29,
30,
34]. Key information used for estimating ecological habitat service values includes an understanding of baseline conditions and the development of appropriate indicator metrics. Baseline conditions are important in that they serve as the point from which lost and gained services are compared. Reducing uncertainty regarding baseline service flows will provide a more accurate estimate of a projects flow of service benefits and potential service losses over time.
Ecological service losses and gains can be measured using a variety of indicator metrics and units. In many cases, the quantification of changes in ecological service flows is based upon selecting or developing an indicator (can be one or more metrics) that acts as a surrogate to represent the ecological service flows provided over time by the habitat and expressing the changes in services (for that indicator) under the different alternatives as a percentage change from a baseline or reference condition. Ecosystem studies, literature-based information, or a combination thereof, can be used to estimate how ecosystem services may change given an action. The metrics or indicators are selected based upon the site, habitats, and the scientific knowledge of those conducting the evaluation. The metric(s) are typically incorporated into a curve that represents the loss or gain of services over time.
For example, changes in one metric (e.g., fish density) may be selected as a surrogate to represent the loss or gain of ecological services provided by a riverine system. As presented in
Figure 9, fish density was used to evaluate the service losses associated with an oil spill into a riverine system. Fish density data were collected in the river after the oil spill for two years to define the initial shape of the recovery curve. Subsequently, available literature data were used to support the development of the completed ecological recovery trajectory (
Figure 9). In this example, fish density was evaluated in relation to the baseline condition and represented as a percent loss from the baseline condition. The estimated loss of ecosystem service value, given the 100 hectare impact area, was about 142 SHYs. In other cases, multiple metrics (e.g., invertebrate density, fish density, vegetation stem density) may be evaluated to represent the service flows from a habitat (
Figure 10). When multiple metrics are used, their values can be weighted relative to one another to arrive at an overall service loss or gain curve. A conceptual example of a weighted curve is presented in
Figure 11. In a manner to that presented in
Figure 10,
Figure 11, one or more indicator metrics can be selected to represent the gain in ecological service flows from an action that provides ecological benefits above the baseline condition. The ability to select indicator metrics on a case by case basis is important as it increases the flexibility of the approach for various sites and situations.
It should be noted that the inclusion of biodiversity metrics from which to evaluate ecosystem service flows and human well-being is a key consideration in evaluating green practices. Biodiversity is a term that expresses the variability and variety among the living world and can include consideration of species biodiversity, genetic diversity, and ecosystem diversity. Biodiversity provides ecosystem services directly and indirectly to human well-being. The protection of biodiversity is therefore important in the evaluation of a sustainable program. The NESA approach focuses on evaluating marginal changes in ecosystem services on a project specific basis. One of the most common measures of biodiversity that can be incorporated into the HEA methodology to support the estimation of ecological service flows is the species richness metric. This metric can serve as a key surrogate to represent ecological service flows and can be weighted higher, if desired, in relation to other metrics if multiple metrics are selected. As the scale of the project area becomes larger, for example on a large landscape basis such as a watershed, region or country, biodiversity metrics may serve as a primary indicator in developing the shapes of the curves presented in
Figure 4,
Figure 5,
Figure 6. The level to which biodiversity metrics are included within the NESA approach would be a site specific and case by case consideration depending upon the project goals and the scale of the project.
Figure 9.
Fish density as the indicator metric used to evaluate the ecological service losses associated with an oil spill into a riverine system.
Figure 9.
Fish density as the indicator metric used to evaluate the ecological service losses associated with an oil spill into a riverine system.
Figure 10.
Use of multiple indicator metrics to represent the flow of ecological service losses associated with an oil spill into a riverine habitat.
Figure 10.
Use of multiple indicator metrics to represent the flow of ecological service losses associated with an oil spill into a riverine habitat.
Figure 11.
Weighted ecological service loss curve based upon the multiple indicator metrics presented in
Figure 10.
Figure 11.
Weighted ecological service loss curve based upon the multiple indicator metrics presented in
Figure 10.
Aside from a base year, the calculations involve a discount rate that allows for the gains and losses to be evaluated from a net present value (NPV) standpoint. Within the HEA methodology, calculations of ecological service losses and gains associated with a project are computed over time. A discount rate can be applied to the HEA units (e.g., SHY, SAY) since they represent time accumulated service flows [
21,
26,
27,
28,
29,
30,
33,
34]. In these cases, the units are typically displayed as a dSHY or a dSAY which represent ecological habitat values. The discount rate is the rate at which the public is indifferent to consuming goods now or sometime in the future. In evaluating ecological habitat service flows, applications of the HEA methodology must account for the absolute difference in the flow of services from the ecological resources resulting from green practices and how those services are distributed over time. In addition, for any given green practice, it is likely that the ecological habitat service loss or gain may not always be constant over time (
Figure 4,
Figure 5,
Figure 6). An assessment of the shape of the various ecological habitat service flow curves over time is necessary. Quantifying the ecological service value of a green practice can be conducted prior to project implementation through the use of projected metrics based on scientific data and professional judgment. This approach would allow green practice alternatives to be compared in a manner similar to how compensatory restoration project alternatives are compared within the NRDA process.
Within a NESA, HEA is a key method used to quantify changes in ecological habitat service flows. HEA has been upheld in U.S. federal court as an appropriate method to evaluate habitat service losses and habitat service gains associated with restoration [
35,
36]. Parameter assumptions can be made in a relatively short time frame to develop the shapes of potential ecological habitat service benefit or service loss curves. In other cases, detailed longer-term studies can be employed to generate data to support the development of the curves. The HEA methodology has been used to provide a practical and flexible approach to generate sufficient information to support decision-making. Curve development should focus on reducing uncertainty to an acceptable level. In cases where uncertainty exists, conservative assumptions can be incorporated into the development of the curves. Aside from the quantification of ecological habitat services using HEA, there are other methods to quantify changes in direct human use services (
Figure 7) associated with green practices and actions affecting the environment. These are discussed in the following section.
3.1.1.2. Human Use Service Quantification Methods
Ecosystems generate several different types of benefits for humans including benefits that are enjoyed directly through consumption, indirectly through their support and production of directly enjoyed goods and services, and through non-consumptive means such as intrinsic and existence values. Together, the total sum of the different types of benefits generated is the total economic value of ecosystem services. There are several economic models that are capable of estimating different components of total economic value. Models that are specifically capable of estimating values of public goods and services include hedonic pricing, stated preference methods, and the travel cost model [
37]. Each method measures a different type of public value. For example, non-consumptive values can be estimated using stated preference methods whereas direct-use values can be estimated using the travel cost method.
Hedonic models use observed price differentials in housing markets to assess the marginal implicit price that local environmental attributes add to housing prices [
37]. Housing prices are a function of housing characteristics, neighborhood qualities, and environmental aspects. When the various determinants of housing prices are controlled for such as number of bedrooms and size, such purchases reveal the values people hold for environmental goods [
38]. Hedonic pricing models are commonly used to assess the impact of pollution and proximity to open spaces [
38,
39].
Stated preference methods are capable of measuring values that are not observed by the choices people make [
37]. These methods use survey instruments and hypothetical situations to assess non-use values such as preservation and conservation, bequest values, and intrinsic value of open space and species. Survey instruments are designed to walk respondents through a situation in which the quality or quantity of an ecosystem service will change. The goal of the exercise is to obtain information on individual’s “willingness-to-pay” for the environmental amenity where “willingness-to-pay” is defined as the maximum sum of money the individual would be willing to pay rather than do without an increase in the ecosystem service [
37].
The travel cost method is able to measure the direct use of nature, whereas, hedonic pricing and stated preference methods measure indirect and non-consumptive values. The value derived from the direct use of nature is observed through the choices people make regarding the use of natural resources. The travel cost model measures the direct use value of nature by examining how much individuals expend (
i.e., time and money) to experience the natural environment in a variety of recreational, educational, and cultural ways. The value of these services is estimated by assessing the net gain to society that is derived from the resource base [
37]. This net economic value, known as the consumer surplus, is the additional value an individual holds for recreational use of nature above the cost required to enjoy the outdoor experience [
37]. When the resource quality or quantity changes, the additional value above expenditures will increase or decrease. Therefore, changes in the availability, access, or quality of the recreational experience can be estimated by examining the difference in the baseline value and the resulting consumer surplus associated with the project or management plan.
Conducting primary economic studies may often be infeasible given the cost and time required to collect and analyze the data. Therefore, many analysts estimate publically valued ecosystem services using benefit transfer methods. Benefit transfer refers to methodologies that use knowledge gained from past studies regarding the value of similar services at comparable locations and employs this knowledge at a new location. There are several ways to transfer values including unit-value or mean value transfer, a benefit function transfer, and a meta-analytic transfer [
40]. Transferred values are adjusted for population, income, and other factors in order to obtain an estimate of the value at the project location [
41].
Valuation of the direct use of nature using benefit transfer methods requires quantification of recreational trips and the net change in value resulting from the project. Quantification of trip equivalents should include all the different activities available, the number of users for the different activities, and other information such as variability and seasonal use. Estimation should be adjusted for any anticipation in the change in the number of trips taken per season due to project implementation. Monetization of direct use can be estimated by transferring per trip consumer surplus values estimated at other locations for each activity. Once direct human use services have been quantified and monetized, the flow of service value is projected through time. The stream of value can be discounted and presented in net present value terms. If the stream of value is positive, this indicates the net change in service level is above baseline. A conceptual example that demonstrates the gain in human use recreational value associated with the enhancement of a green space is depicted in
Figure 12. The enhancement could entail improvements to habitat quality and/or improved access to the habitat.