The notion of “sustainable development” has permeated significant parts of policy discourse about the environment. This reflects a number of (related) concerns including the development path that the broader economy is on and specifically the way in which (changes in) natural wealth affects this path. It is important that CBA speaks to those concerns especially as policy and investment projects have the potential to shift a development path (perhaps because of non-marginal actions or the cumulative effect of smaller decisions). There are a few implications of this but one of the most prominent (as well as far-reaching) is to circumscribe CBA by having it live within sustainability constraints, perhaps based on ecological criteria. This places greater emphasis on a single appraisal within the context of a portfolio of policies or projects. That is, the constraint is that this portfolio, on balance, maintains the ecological status quo with practical applications of this approach including biodiversity offsetting. This raises important issues. On the one hand, there is a benefit to avoiding untoward and irreversible damage to (possibly) critical resources. On the other hand, there are opportunity costs to applying the shadow projects approach that need still to be considered.
Cost-Benefit Analysis and the Environment
Chapter 12. Sustainability and natural capital
Abstract
12.1. Introduction
How far does current CBA practice impart information about the sustainability of policies or projects being evaluated? Clearly, any answer to this question is, in large part, dependent on defining what is meant by the term itself: “sustainability”. For example, it might refer narrowly to the internal sustainability of the project itself perhaps because of financing risks and budget constraints. Alternatively, it might refer much more broadly to a whole range of external factors – economic, social or environmental – which could be influenced by an investment project or policy decision. An illustration of the possible breadth of such factors is the 2015 UN Sustainable Development Goals: 17 high-level development objectives with more than 160 sub-objectives.
This breadth could be viewed as implying the need for a multi-criteria approach (see Ch. 18). In some transport applications, for example, the sustainability challenge for appraisal has been interpreted in this way. In this chapter, however, a more specific conceptualisation of sustainability is adopted; one that is drawn from economics albeit with broader interdisciplinary implications (see, for example, Arrow et al., 2012; CGDD, 2015; Hamilton and Atkinson, 2006; Helm, 2015). Its distinguishing features are sustainability as a term articulating concern about future generations (i.e. intergenerational equity) and, as a way of addressing this concern, a focus on what is happening to wealth (e.g. future well-being prospects and assets in an economy, particularly natural capital) as a result of proposals for investment projects and policies.
What then is distinctive about this “sustainability economics” in informing economic appraisal? At the very least, it draws together a number of compelling critiques of CBA. One is scepticism of the Kaldor-Hicks criterion combined with the remedial prescription that interventions which harm certain groups (e.g. future generations) should be accompanied by actual compensation. A further concern is that the “marginal” or “incrementalist” approach of much of practical CBA does not consider the sustainability of the broader “system” (Helm, 2015). This might refer to prospects along the development path of an economic system. Such considerations have been a feature of climate economics too (Chapter 14). The emphasis in “sustainability economics”, however, is on the way in which a development path is influenced by the culmination of (policy- and project-induced) changes in a whole range of natural systems, characterised commonly as “natural capital”.
Dealing with these concerns still leads to a number of distinct possible paths. Indeed, one response might be that there is little wrong with the way with which CBA is conducted and that existing knowledge is evolving in ways that cover a number of the challenges set out by those concerned about sustainability. This is not an indefensible position. A lot of CBA certainly does deal with relevant challenges: environmental valuation, discount rate selection, and decision-making under (future) uncertainty are all relevant examples of this.
For example, given that a lot of the concern about sustainable development is based on judgements about distributional outcomes this reinforces a need to report how costs and benefits are distributed over time. This call is not new (e.g. IPCC, 1996) but has been recently reiterated in Day and Maddison (2015) (see Chapter 10). There is a need to improve the reach and accuracy of methods – including techniques of environmental valuation – for measuring (changes in) natural capital, rather than simply measuring flows of current services. Reconciling CBA with sustainability concerns may also force explicit thinking about how much of nature should be put beyond straightforward cost-benefit thinking, and thus what this implies for how CBA is conducted.
12.2. What is sustainability?
While the question as to “what is sustainability or ‘sustainable development’?” is often posed as one that cannot be answered – or at least is a question which has many, perhaps contradictory, answers – this is little help in terms of clarifying how practitioners might better integrate sustainability thinking into CBA. So if, for example, it is defined extremely broadly (i.e. covering “all” aspects of the development process and multidimensional outcomes) as in many existing national sustainable development strategies or the United Nations’ SDGs – then potentially the challenge is huge. That is, it is arguable that CBA cannot possibly cover the breadth implied by these high-level frameworks. In such cases, a logical reaction might be to incorporate these concerns in appraisal by using additional assessment tools in making a decision (see Chapter 18).
If instead sustainable development is conceptualised less broadly – perhaps in terms of an economic definition of non-declining (per capita) human well-being over time and, in turn, in terms of how wealth or assets in an economy are managed to achieve that end – then the challenge becomes more tractable, or at least implications can be understood more clearly. This economic approach to sustainability does not have all the answers. However, it provides a useful starting point for understanding a coherent core of the challenge, posed by the sustainability debate. In this way, extensions to that understanding can then be considered.
On this view, future (development) prospects depend on the wealth that an economy has. Projects and policies represent one way in which these prospects are affected. The impact might be large or small but the projects and policies being appraised in a CBA can be interpreted as shifting the development path of an economy over time (e.g. Arrow et al., 2003). These interventions often do this explicitly, perhaps by investing in assets in the economy. Prominent examples traditionally would include physical infrastructure projects in the transport or communications sectors, investment in the public education sector or influencing the health of the nation by spending on better water treatment and sanitation services. However, the impact may also be implicit by creating (potentially) investable resources because an intervention generates net benefits and so entails better prospects for future consumption or well-being.
How projects and policies affects the natural or physical environment is a further, but related, consideration. This natural wealth matters for development because it is itself a determinant of future well-being by providing flows of goods and services that ultimately provide people with benefits that they value. For this reason, this is increasingly referred to under the broad heading: natural capital. What then is “natural capital”? While useful as an umbrella term, a more specific definition is needed as ever to be of use to cost-benefit practitioners. The UK Natural Capital Committee, for example, defines natural capital as: “… the elements of nature that produce value to people (directly and indirectly), such as the stock of forests, rivers, land, minerals and oceans, as well as the natural processes and functions that underpin their operation” (NCC, 2013, pX). What links this diverse and large range of naturally occurring resources is that these are stocks. Further properties of these stocks are also relevant to distinguish too, not least because these might imply different recommendations for consideration in a CBA (or social decision-making, more generally).
First, some of these assets are non-renewable. These stocks are fixed in physical extent (although there might be uncertainty about what this extent is), are non-living resources and so are exhaustible. Sub-soil assets, such as conventional oil and gas, are typical examples. Second, some of these assets are renewable. These are living resources or resources which regenerate (perhaps because of natural growth or some other underlying natural regenerative process). A forest or a fishery are well-known examples for CBA practitioners. Ecological or ecosystem assets are a further category that has gained increasing attention in policy debates (see below). These renewal properties lead to further questions about resource management. That is, renewable resources can be sustained at some level even if used, in contrast to non-renewable resources.
What then does this imply, if anything, for cost-benefit practitioners? Of particular interest perhaps are those implications when the investment project or policy being appraised will result in the loss of natural capital. For example, for non-renewable natural capital (such as a mining project or depletion policy) there is a corresponding question about what supplementary actions (if any) must accompany this activity in order to compensate for this loss of asset value. If the asset loss involves some renewable resource – as in the case of land-use change which results in degraded or destroyed ecosystems – then sustainability might imply asking whether this loss can be compensated by building up other assets generally, or whether it requires specific offsetting investment in the renewable asset itself.
12.3. CBA and “weak sustainability”
Building this picture of the challenge that concern about sustainability poses for economic appraisal might start with taking stock of what conventional CBA currently does. In this spirit, Stavins et al. (2003) reflect on what the cost-benefit criterion, and compensation tests, signals and how this might be interpreted in the light of concerns about intergenerational equity. This starts then with the basic Kaldor-Hicks criterion: choose proposals with (maximum) positive net present value, such that winners potentially can compensate losers and still remain better off than without the project or policy change. Where “winners” are those in the current generation and “losers” are future generations, the argument goes that at least this signals potentially available economic resources to address concerns about intergenerational equity arising from these interventions.
While in one sense all this does is restate that standard cost-benefit perspective, it usefully draws a link between efficiency and equity over time and between generations. It prompts, for example, immediate questions about not only whether there are mechanisms to straightforwardly facilitate compensation between generations and whether potential compensation really does settle unease about intergenerational equity. One study then that takes this a bit further is an earlier contribution by Farrow (1998) who argues that this compensation – in these sort of cases – should be actual rather than potential. This starts by asking explicitly whether a proposal imposes net losses on a particular group: i.e. future generations? If so, compensation actually offered to that particular group, must be at least as great as these net losses. Linking this to what might be happening to natural capital as a result of current decisions, this compensation must be at least as great as any loss in asset value that the proposal causes. The idea here then is that such losses in asset value decrease well-being by diminishing future consumption possibilities. One way of, at least, maintaining these possibilities is by investing in other assets to offset these current losses.
Such ideas are as the genesis for a practical recommendation in, for example, Day and Maddison (2015). They lead with the point that CBA provides useful information about positive net benefits of proposed actions that signal potentially investable economic resources. If a proposal also results in costs which amount to losses in asset values (perhaps because natural capital is depleted or degraded) then the corresponding gains can be used to invest to offset this and, as such, keep capital constant. As for whether this investment is forthcoming in reality, this is a decision in the hands of decision-makers of course. But if decision-makers believe in using CBA to inform policy choices along with a commitment to broader sustainable development then an economic appraisal at least signals the economic resources made to realise this latter obligation.
For practical purposes, this suggests two bits of policy-relevant information should be co-joined. The first is the standard one from a cost-benefit perspective: recommend projects or policies which pass a standard cost-benefit test. The second is to check whether the totality of assets in the economy is at least being held constant. This latter element could be assessed by looking at an indicator of how these assets are changing. As discussed below, such a metric is increasingly referred to as adjusted net – or “genuine” – saving. What this indicates is the extent to which net assets are being accumulated in the economy and, importantly, this includes changes in natural capital. Whether “genuine saving” is positive or otherwise possibly provides this information about the broader sustainability of economies within which CBA is being conducted on specific decisions.
This practical link between changes in total wealth and sustainability was first explored by Pearce and Atkinson (1993). Subsequent growth theoretic literature, including Hamilton and Clemens (1999), Dasgupta and Mäler (2000) and Asheim and Weitzman (2001), has elaborated the theoretical foundations for this approach to measuring sustainability. The basic insight is contained, however, in Hartwick (1977) which showed that future consumption can be sustained when exhaustible resources are extracted if other investments offset the value of resource depletion. By definition, an investment in a mine that enables the extraction of a valuable but finite deposit of some natural resource is financing an unsustainable activity. That is, mining can continue only up to point that the resource is exhausted (either physically or economically). The broader implication for sustainability is another matter. Much depends on whether or not the proceeds from mining the resource are invested in an alternative (productive) asset. If the proceeds of mining are ploughed back into new and productive projects then development can be sustained.
Formally, Hartwick (1977) showed that important earlier insights about this problem (such as Solow, 1974) imply a savings rule where some portion of the revenues (specifically, resource rents) from the depletion of an exhaustible resource is invested in alternative assets. These alternatives are usually thought of as produced assets, but it could also be human capital. In this way, development – defined as the constancy of a consumption path – can be sustained in perpetuity despite the dependency on a finite (exhaustible) natural asset. The way to ensure this is to keep total net savings across all types of capital – or “genuine saving”, to use the term coined by Hamilton (1994) for this concept – at or above zero. Solow (1986), in turn, showed how following this guidance implies a constant capital rule. This has become the bedrock of modern thinking about sustainability economics.
A general form of this “Hartwick rule” is a more recent development but has an important property: it is consistent with a possible consumption path which is increasing. Hence, this can accommodate a policy regime where economic growth is the objective rather than just the constancy of economic development along a path. Hamilton and Hartwick (2006) identify a relationship between (the change in) consumption and net saving. This generalised Hartwick rule posits that positive genuine saving can lead to growth in consumption along the development path. A key condition of this is that genuine saving does not grow faster than the interest rate (i.e. the returned on produced capital). Hamilton and Withagen (2007) explore the implications of this a bit further, showing that a constant positive genuine saving rate (the share of saving in national income) implies consumption can increase without bound.
Thus, to the extent that a project entails the (net) accumulation of produced assets or human assets then, other things being equal, it contributes to sustainability. In other words, discussion about sustainable development in the context of CBA needs to consider these desirable wealth-increasing properties of many projects and policies. Of course, to the extent that such actions give rise to environmental liabilities or deplete resource stocks then this loss of natural assets decreases sustainability, other things being equal. However, as previously discussed, the net effect is signalled by aggregate indicators such as genuine saving or the change in per capita net wealth.
A practical example of this sort of thinking is discussed in Box 12.1. This illustrates the use of wealth funds to manage the proceeds of non-renewable resource depletion (such as oil and gas). While the literature on “genuine saving” as one element of how policy can realise concern for future generations, less attention has been given to the productivity of investments. Clearly, this latter issue falls squarely within the domain of CBA. Not only can projects, selected by cost-benefit appraisals, increase net wealth but also can further contribute to sustaining development by ensuring that savings are put to the most productive use.
Box 12.1. North Sea oil and sovereign wealth funds
Exhaustible resources and the revenues they generate present two broad problems for macroeconomic management: gross production and tax revenues tend to be large and highly volatile, and the stream of revenues is finite, ending when the resource deposit ceases to be economic. Large flows of resource tax revenues lead to the distinct risk that fiscal policy will be pro-cyclical and hence a source of macroeconomic instability. And the finite nature of the resource revenue stream raises important questions about the sustainability of the macro economy – will well-being fall as the resource is exhausted?
A number of countries are turning to sovereign wealth funds (SWFs) as a way of handling these risks and opportunities including notably Norway which established its fund in the 1990s. Hamilton and Ley (2011) list 12 countries or jurisdictions where resource funds and/or fiscal rules for resource revenues have been implemented. The United Kingdom is an exception here having decided in the late 1970s not to establish a SWF and subsequently never (it appears) revisiting the issue. But given that North Sea revenues reached 9.9% of fiscal revenues and 3.7% of GDP in 1984, with revenues exceeding 1% of GDP from 1979 to 1987, it is fair to ask what are the costs and benefits of that decision.
To explore the likely sacrifice from the establishment of a (hypothetical) sovereign wealth fund in 1975, Atkinson and Hamilton (2016) carry out an ex post cost-benefit analysis of an SWF as a public investment. Resource revenues transferred to the fund become costs from the Treasury viewpoint, while fund payouts to the Treasury are benefits. For simplicity they assume that the petroleum resource was depleted by 2010. Costs of investing in the SWF cease at this point. The fund, however, is assumed to continue paying out an amount equal to its 2010 value (the real return on the portfolio) in perpetuity. Table 12.1 shows the present value of costs and benefits of investing in the SWF, normalised per capita, for different time horizons and discount rate assumptions.
Table 12.1. Costs and benefits of establishing a sovereign wealth fund
Source: Atkinson and Hamilton (2016). The discount rates are derived from Green Book 2003. Levelised costs and benefits are calculated over 1975-2010 or 1990-2010 as appropriate.
To isolate the effects of the oil price bubble and other economic circumstances in the early 1980s, the calculations simulate costs and benefits for two assumptions about the year that investment in the fund commenced, 1975 or 1990. Starting in 1975 yields a larger present value of benefits from the fund, but it also yields a high present value of costs because these costs (payments into the fund) are front-end loaded. This front-end loading is much diminished for the fund simulated to commence in 1990.
The first discount rate assumed, fixed 3.5%, is equal to the Green Book 2003 social discount rate for projects where costs and benefits span 30 years or less. The second discounting scenario uses a discount rate of 3.5% that declines beyond 30 years. This is the Green Book 2003 social discount rate assumption for assessing policies affecting the long term – which is precisely what a SWF is designed to do.
As shown in Table 12.1, the present value of benefits from the SWF exceeds costs by 10% to 69% for a fund starting in 1975. For a fund starting in 1990, the corresponding figures are 11% and 100%. In terms of net annual benefits (levelised benefits minus levelised costs), these vary from GBP 14 to 10 per capita for the fixed discount rate, and GBP 97 to 88 for the declining discount rate – in each case, the 1990 scenario yields the lower net benefit figures. The fixed discount rate results are more sensitive to the choice of discount rate, with a discount rate of 3.92% (the assumed constant real rate of return of the SWF) yielding 0 net benefits for either start year. This is an artifact of the synthetic nominal SWF return that the authors use to simulate fund returns (which is, in turn, based on average real returns on a globally weighted mix of holdings of equities and bonds).
Establishing a SWF would not have been without sacrifice. However, this ex post analysis suggests that if a SWF had been established per Green Book 2003 standards, the net benefits per capita would be positive and moderately large for the assumption of declining discount rates. Importantly, it also potentially generates a sustained source of income from using a finite resource
An interesting development is extensions to settings including renewable natural capital. This leads to at least two important and overarching questions. First, this requires advances in environmental valuation both conceptually and empirically. Secondly, nor are such extensions without controversy. The economic approach to sustainability elaborated above is often labelled under the heading of “weak sustainability”. This is a descriptive label used to distinguish this approach from stronger approaches which emphasise far more of a special place for conserving natural capital in thinking about intergenerational concerns. As such, this label is prescriptive too distinguishing it from approaches which are perhaps a lot less permissive in guiding how cost-benefit practitioners respond to the challenge of sustainability. Both these issues about natural capital valuation and “strong sustainability” are considered below.
12.4. Valuing natural capital
A lot of the existing terminology in environmental valuation has focused on valuing flows: that is, some flow of a benefit arising perhaps from the consumption of a good or service. Of course, policy interventions such as investments in ecosystem protection (or enhancement) typically will boost the flow of these services over time, thereby introducing a dynamic element into any economic analysis. Moreover, when these same ecosystems are perturbed by some change (be it a shift in land use or a degradation in state) the effect on well-being similarly will have an intertemporal dimension (e.g. Mäler, 2008; Dasgupta, 2009). Put this way, what one needs to think about is the underlying ecosystem asset and, in particular, the changes in asset value that occur as a result of human interventions (be these positive or negative, deliberate or otherwise).
Thus, these flows of services can be viewed as the flows of “production” that are supplied by underlying assets or “natural capital” (e.g. forestland, wetland and so on). Recalling the discussion in the previous section, what needs to be assessed here is the potential change in the future prospects given what is happening to this natural capital now. In doing so, this might throw light on whether natural capital use and economic development paths more generally are sustainable or not.
There remains a strong connection to CBA principles in this work. One example is those contributions which seek to assess economic sustainability when there is land-use change such as deforestation as loss of other natural habitats such as mangroves. The basic unit here – following Hamilton and Atkinson (2006) and Barbier (2009) and earlier contributions, particularly Hartwick (1992) – is land. That is, land under a particular use has a distinct asset value. When land-use is changed – as happens when forestland is cleared – this can be viewed from the perspective of CBA; that is, is the change net beneficial? Additionally, there is a corresponding implication for how wealth is changing in the economy that should be accounted for.
In the case of deforestation, decrease in the value of forestland leads also an increase in the value of agricultural land assets. Put another way, what has happened here is a change in composition of the broader portfolio of land assets. For example, if one ecosystem service provided by woodland is climate regulation (via carbon sequestration and storage services), increasing the amount of woodland will increase the provision of these services. But there is likely also to be some loss in the climate regulation services provided by agricultural land and, ideally, these services that are lost also need to be recorded somewhere. Forestland is also an asset providing multiple benefits and it is important that as many of these are accounted for as possible. There are further measurement issues. Clearly, other services that change as a result of the land-use switch need to be accounted for. The broader balance sheet, for example, will reflect the loss in agricultural output and so on. There are also presumably conversion costs associated with changing land use and those investment costs should also be accounted for.
A major element to incorporating better measures of (changing) natural capital into CBA is an extension of valuation to this domain. Progress on environmental valuation – and particularly in the realm of ecosystem (service) valuation – offers some encouragement here. But important debates remain about this progress. And regardless of whether one views the glass as half-full or half-empty in interpreting those debates, addressing this valuation challenge remains work-in-progress. For example, it is clear that a great many development projects have an impact on biodiversity; i.e. by changing land-use and natural habitats. However, it is far from clear that even state-of-the-art appraisal can provide an adequate assessment of the value of this loss. As such the suspicion might be that natural capital valuation might be painting a sufficiently full picture about what happens when investment projects and policies affect natural assets such as ecosystems (and underlying assets such as biodiversity). Judgements about the ability of practitioners to rise to that challenge may require a more circumspect role for CBA in this regard (see next section).
Where valuation is possible, the problem seems analytically more straightforward at least on the face of it. That is, the value of natural capital can be viewed as equal to the capitalised value of flows of future services. Of course, for many categories of (renewable) natural capital such as ecosystems, the prices of the resulting flow of services are not observed and so neither are the prices of ecosystem assets. However, as a number of chapters in this volume illustrate, considerable progress in environmental valuation at least allows an ever more complete response to this challenge (see also Chapter 13). As Box 12.2 illustrate, some practitioners have sought to distil this growing evidence base into truly ambitious estimates of the aggregate value of changing natural capital.
Box 12.2. The value of aggregate natural capital
The recent emphasis on large-scale ecosystem assessments – such as the OECD Environmental Outlooks (e.g. OECD, 2012), TEEB and the UK NEA – indicates some interest in searching for clues about the overall scale – in economic terms – of what has been lost (and what is likely to be lost in the future) as a result of the continued destruction of the natural world. While this is not a substitute for more detailed policy analysis, knowledge about these trends might be important for framing policy thinking. In addition, such information might throw light on whether ecosystem and biodiversity decline is a development problem as, for example, Stern (2007) demonstrated in the case of climate change.
There are, however, clear signs of growing interest in this question. An example of this is the linkages being made between (recent and on-going) ecosystem assessments and efforts to understand the way in which changes in natural wealth influence the sustainability of development through greening of national accounts (see, for example, World Bank, 2010; Arrow et al., 2011). The on-going World Bank led consortium “WAVES” project (Global Partnership for Wealth Accounting for the Value of Ecosystem Services) represents a practical application of this work to a number of proposed countries. This has clear relevance to the question at the heart of the economic approach to sustainability: i.e. is enough being saved for the future?
Some studies have sought to explore these issues but do so by calculating losses in natural assets likely to occur according to possible policy scenarios (and hence in principle ask a more defensible question than that about the totality of the current service flow). Hussain et al. (2012) estimates the losses arising from recent past and projected future loss of the world’s aquatic ecosystems (specifically wetlands, mangrove and coral reefs). The present value of this loss over the period 2000 to 2050 (using a discount rate of 4%) is reckoned in excess of USD 2 trillion (in 2007 USD) (with two-thirds of this accounted for by wetlands). The annualised value of this total change is just under USD 100 billion (that is, the value of the loss of these ecosystem assets each year is estimated to be of this magnitude) which, e.g. in 2007, was just 0.2% of global gross income. Chiabai et al. (2011) conclude not entirely dissimilarly for the case of the loss of global forests over the same time period.
Needless to say, such global estimates of ecosystem loss require some heroic assumptions and generalisations (with the same being true of efforts elsewhere to value the global impacts of climate change). Indeed, for some critics, a search for a global value is a flawed project because of this. However, taking these findings at face value it appears that knowing the global magnitude of ecosystem losses might not add significantly to empirical discussion. So while it is entirely possible that these analyses are missing something possibly both large and critically important, a tentative conclusion is that the pragmatic demand for more highly aggregated indicators of trends and concerns about validity both point away from an emphasis on global trends.
Greater practical significance, however, is to be found at the regional or country level. In the case of forests, for Brazil, estimated losses in natural wealth are found by Chiabai et al. (2011) to be substantial (as a percentage of the country’s gross national income or GNI). Hussain et al. (2012) find that for aquatic ecosystems, for the South Asia region and for Indonesia, however, these annual losses in natural wealth were respectively 1.7% and 4.0% of GNI (in 2007).
These are magnitudes worth knowing more about. This would necessitate still close scrutiny about the robustness of such estimates. The basic problem of accounting for the value of ecosystems can be put simply. It entails identifying a price or (unit) value and a quantity of (some change in) the provision of e.g. ecosystem service (Boyd and Banzhof, 2007). An immediate challenge, however, lies in identifying the likely limits on how the available empirical record on ecosystem “prices” and “quantities” can be pulled and stretched over the assorted ecosystem areas needed to make robust aggregate generalisations. The issue of spatial variability is central here. This includes properly accounting for variation in the supply characteristics – the type and extent of functions – of ecosystems as well as demand characteristics – of the human population that consumes services that these functions give rise to. All this requires relatively sophisticated mapping and is demanding in information terms. However, it might be that at this national level (or sub-national levels) that these issues become a little more tractable (see, for example, Kaveira et al., 2011).
Of course, much of what is currently termed “ecosystem services” already may be reflected in the national accounts. This is a point made in World Bank (2010). Examples of this might include the natural pollination services that (in effect) are capitalised in the value of agricultural land or the recreational opportunities that are (implicitly and in part) provided by natural areas. On this view, ecosystems support market activity in a number of important (but indirect) ways and the accounting challenge is to correctly re-attribute the service value to the (ecosystem) asset which gave rise to it (Nordhaus, 2006).
To the extent that ecosystem services are missing entirely from the accounts then clearly these will be neglected in any accounting approach that only looks at re-attribution.1 For example, many types of cultural services might fall out of the reckoning in this way. However, as a starting point, an emphasis on identifying what is already (somewhere) in the accounts has merit. In particular, given the traditional opposition by the national accountants to non-market valuation in relation to the accounts (Hecht, 2008; de Haan and van de Ven, 2007), this starting point has a strategic benefit.
1. Vanoli (2015) proposes an approach that integrates a measurement of the deterioration of ecosystems in the national accounts.
The challenge for natural capital valuation is not confined to progress in techniques to measure environmental values. Recommendations about discount rates are important too in capitalising current services in the appraisal of impacts which – given the potential longevity of renewable natural capital – arise in the distant future. So debates about the size of the discount rate – and the term structure of the resulting discount factors – are important here too if the future consequences of current actions are not simply assumed away by the “tyranny of discounting” (see Chapter 8).
Valuing natural capital, however, is likely to involve more than simply capitalising flows of current services. As shown in Fenichel and Abbott (2014), accounting for the value of an ecosystem asset (or renewable natural capital more broadly) requires estimation of a range of parameters, of which the value of the flow of ecosystem services is just one ingredient. First, when the asset is renewable (or regenerates), the ongoing resource productivity must be considered in discounting the (future) value of the asset. Second, there is a capital or holding gain, which Irwin et al. (2016) term as a “scarcity effect” arising from holding the last or marginal unit of the asset.
That is, asset price, 
, where V is the value of the marginal unit (current) service flow from the asset and r is the discount rate. One additional term, in the denominator, is 
, which refers to (net) resource productivity and is used therefore to calculate an effective discount rate. A further term, in the numerator, 
refers to the “scarcity effect” of holding the last unit of the asset. This suggests then (depending on the magnitudes of
and 
) that the simple approach – i.e. capitalisation of current services – is missing something important. As such this has potentially important implications for accounting for natural capital asset values, as well as degradation or enhancement of these assets.
Fenichel et al. (2016) apply this conceptual framework to the challenge of valuing groundwater as an asset in rural Kansas in the USA, over the period 1996 to 2005. The stock of groundwater – the amount of water held in an aquifer of a given size – is defined here as the product of the thickness of the saturated zone (most comprised of rock) and an estimate of its yield. During their study period, the authors find that groundwater was being depleted, in physical terms, at a rate of 0.4% per year. However, this underestimates the corresponding annual change in the economic value of this stock. The reason for this is that as the stock declines, its marginal value increases because of a “scarcity effect”. As an illustration of this the authors show that when groundwater is scarce the monetary value of a marginal unit – defined in terms of its value to agriculture – is roughly twice as high as when groundwater is abundant (i.e. around when it is about ten times as plentiful in physical terms). Accounting properly for the economic depreciation of groundwater in Kansas must consider this schedule of prices associated with different degrees of abundance. Fenichel et al. show that this matters empirically: the economic value of the loss in groundwater stocks is 6.5% per year over the study period.
Another type of issue is what to do when environmental valuation of a stock needs to take into account values which are held by people far into the future. Clearly, one cannot possibly know exactly what future preferences will be – that is, what future people will value – beyond those things one can feel confident will continue to be required for survival or basic functioning. The usual practical response to this uncertainty is to assume that future people have the same preferences as those living in the present but to uplift these values to take into account the likely effects of changing (i.e. growing) per capita income. Less common is taking account of the likely path of natural assets. That is, if it is thought that a natural asset will be more scarce in the future, then it is plausible that the (marginal) value that will be placed on future losses of services from this asset will be higher (than now).
A further (but related) question is what happens to such values when the underlying asset is difficult to replace. In the example above from Fenichel et al. (2015), one explanation for the scarcity effect estimate for groundwater is the characteristics of resources, which tends to be localised with limited substitutability. Gerlagh and van der Zwann (2002) consider the general case where these substitution possibilities are a function of the asset stock itself. That is, when a resource such as an ecosystem is relatively abundant, losses in that asset “do not matter” in the sense that this source of well-being could be easily replaced with something else and people essentially would be no worse off. However, after some threshold, substitution possibilities diminish rapidly. In other words, continued loss of the natural asset – beyond this critical point – increasingly cannot be compensated and, on the contrary, increases the prospect of ever higher and higher adverse impacts on future well-being as the resource continues to be depleted.1 The implications of limited substitutability can be complex. Traeger (2011) shows that this affects the magnitude and term structure of the social discount rate (see Chapter 8). But the basic point remains that a lack of substitution possibilities should translate into a correspondingly higher (marginal) value to assign when natural capital is destroyed or degraded.
These critical issues about substitutability are also explored in contributions by Hoel and Sterner (2007) and Sterner and Persson (2008). Both of these papers show how the value (or shadow price) of a scarce environmental amenity might increase over time. A key parameter here reflects the ease (or difficulty) with which particular natural assets can be replaced: i.e. the “elasticity of substitution”. The higher the value of this elasticity (reflecting the greater difficulties of replacing a natural asset with another type of wealth), the faster is the increase of the asset price of an environmental amenity as the natural asset (giving rise to that amenity) becomes more and more scarce.
Using these insights for practical analysis requires that a number of assumptions must be made: most notably, about the elasticity of substitution. Even so, there could be substantial problems such as evaluating future prospects using “sustainability prices” – that is, prices which are consistent with realising a sustainable path (which is a different point to the matter of correcting prices for current market, and other, failures). The point here is that there might be some sustainability problem that appraisal at the project level or aggregate level, for that matter, will not pick up. In principle, a conventional CBA might be able to address these issues. However, in practice this might be fraught with difficulties and may necessitate a rather different treatment of these same concerns within an economic appraisal.
12.5. CBA and (strong) sustainability
This issue of limited (or a lack of) substitutability between natural capital and other assets has important implications for rules for the sustainability of development as well as how CBA might be conducted or interpreted. Part of the challenge is the analysis of what happens when complex assets, such as renewable natural capital, change as a result of policy interventions. One example of this is the concept of “ecological resilience” that might characterise ecological capital (Mäler et al., 2009; Mäler, 2008): the ability of an ecosystem to withstand stresses and shocks (and to continue to provide services).2 Walker et al. (2010) looks at the value of this resilience to agriculture in South-East Australia of maintaining a saline free water table (mainly through farmers cutting down trees to expand agriculture). Here agricultural expansion represents a driver depleting the stock of non-salinated soils (measured as the depth of soils for which saline intrusion is not a problem). As this depletion driver is increased so the stock of ecological resilience falls. The depleting process itself may generate benefits (here agricultural produce) and so there is a trade-off to be assessed between the benefits of depletion and the fact that losses of resilience may need to be reversed if stocks fall below some threshold level. Breaching the threshold, however, leads to likely irreversible losses in agricultural productivity (because of salinated soils) so this resilience has a distinct value which should incorporated in economic appraisal of actions which move this system towards or away from threshold.
Another problem arises from non-linearity. A cost-benefit analysis that fails to account for thresholds, for example, might recommend the conversion of part of an ecosystem, or other for more direct human use. But the assumption might still be that conversion of this part of the ecosystem would not affect the remaining services provided by the rest of the ecosystem. Non-linearity means that this assumption could be suspect. The real difficulty here arises from interdependencies between the various services provided by the ecosystem. In terms of valuation this means that the economic value of any one service may depend on its relationship to the other services. Valuation concerns changes in the ecosystem, and this is itself dependent on how everything changes, not just the service that the practitioner might want to focus on.
This is, incidentally, another reason why estimating “total” value is not feasible – as one, say, decreases the ecosystem dramatically, everything will change. The critical point here though is the ecological area is a “system”. Ecosystems have interactive processes, a variable potential to adapt to exogenous change, and the relevant changes are often non-linear (Arrow et al., 2000). If so, then from a policy perspective it should be managed as such and, in turn, this might shape how CBA is used to inform decisions over management options. Importantly, it is not clear that “bottom up” (marginal) approaches whereby each type of service is valued separately and then the values are added to get some idea of the total value of the ecosystem, are capturing the “whole” value of the ecosystem. Put another way, the value of the system as a whole may be more than the value of the sum of its parts. The bottom-up valuation procedure could therefore be misleading. A small economic value for one service might suggest it could be dispensed with, yet its removal could reverberate on the other services through complex changes within the ecosystem.
One further problem is that there is both uncertainty about the nature of the services themselves and, even more so, about their interactions. While many agree that natural capital such as ecosystem assets are characterised by thresholds, there is less certainty about what these thresholds are, especially for the practical purpose of taking account of these in policy formulation tools such as CBA. So the example above for the case of saline-free agricultural land in South East Australia is an exception rather than the rule of what is known empirically. Similarly, while non-linearities may mean that the consequences of losing even modest amounts of ecosystem could be large (as threshold are approached or breached or where there are interactions with other parts of the system).
So converting a natural system may therefore produce unanticipated and adverse effects, which could be irreversible. Efforts at valuation of such changes in a CBA remain important. But it is important to recognise that these are unlikely to provide much information about the scale of “tolerable” change. Moreover, if decisions are made and they turn out to be extremely costly, little can be done to reverse them. Losses of natural capital assets can combine several features: a potential large “scale” effect; irreversibility; and, uncertainty. The argument is that these “strong sustainability” characteristics combine to justify a presumption that natural capital (and its components) should be conserved. The implication is that this view necessitates also a rather different approach to using CBA to make investment and policy decisions.
12.6. Cost-benefit analysis and precaution
Economists have long known that this combination argues strongly for a “precautionary” approach to making decisions (e.g. Dasgupta, 1982). Chapters 9and 10 observed that there are two ways in which to conduct CBA. The first approach – the one that is most commonly used – operates either in a world of low uncertainty or in a context of uncertainty where the appropriate decision might be made in terms of expected values. The second takes more account of uncertainty and also takes explicit account of irreversibility, either because funds committed cannot be “uncommitted” or because other effects of the policy cannot be reversed (or both). This was described as the “real options” approach to CBA.
On the real options approach, considerable attention would be paid to the opportunities for learning, and thus reducing uncertainty, by delaying irreversible decisions. It seems clear that the many aspects of the issue of natural capital (particularly ecosystem change) fits the real options approach: there is uncertainty, irreversibility and a major chance to learn through scientific progress in understanding better what natural assets do and how they behave. It is in this sense that real options gives rigorous content to a notion like “the precautionary principle”. Note that, on this interpretation of the precautionary principle, there would be far more caution about losing ecosystems, but benefits and costs would still be traded off.
Another contender for a precautionary approach would be the “safe minimum standard” (SMS) (Ciriacy-Wantrup, 1968; Bishop, 1978 and Randall, 2014). On this approach, natural capital conversion or loss would not be countenanced unless the opportunity costs – i.e. the value of the forgone “development” – were intolerably high. What the SMS approach does is to reverse the onus of proof, away from assuming that development is justified unless the costs to the environment are shown to be very high, to a presumption that conservation of natural capital is the right option unless its opportunity costs are very high. But determining what is meant by “intolerable costs” is not easy. The level of “tolerance” might be determined by the political process, by reference to some notional benchmark – perhaps a percentage of GNP, or by a more extreme indicator – e.g. the forgone development causes severe hardship or poverty.
The principle of precaution – for the reasons outlined in the previous section and perhaps operating via ethical reasoning and/or a decision-making framework such as the SMS – suggests, in turn, a strong sustainability principle. It argues that no further degradation or loss of natural capital should be tolerated. Incorporating information about scientific thresholds is then one way in which these sustainability constraints can be envisaged. Uncertainty about the location of these thresholds may present a challenge, however. There also remains debate about whether this refers to natural capital in general (assuming some basis for aggregation) or particular classes of natural asset which are critical according to strong sustainability criteria. The tendency, however, is that the latter has been the focus particularly ecosystems such as broad habitat types. The practical implications for CBA are still several. In a very extreme form, this might argue that no existing ecosystem should be degraded. In less extreme form it could argue that any loss has to be offset by the creation of a like asset.
12.6.1. Circumscribing CBA: the example of “biodiversity offsetting”
A number of contributions, beginning with Barbier et al. (1990), have sought to model concern about strong sustainability as a constraint for the purposes of CBA for the reasons outlined earlier in this chapter. While these are largely conceptual contributions, there is also growing practical interest in the application of, for example, resource compensation in assessing real world examples of damage particularly to ecosystems (although the applicability need not be limited to this). The basic approach taken, in theory and in practice, is to recognise that strong sustainability is a concept that is most relevant to the management of a portfolio of projects. That is, for example, imposing a constraint on project selection that each individual project does not damage an ecosystem is arguably too stringent (in the sense that very few projects would presumably yield net benefits yet not damage ecosystems at all).
More flexible proposals for selecting projects and choosing policy options subject to a (strong) sustainability constraint usually advocate that the “net effect” on the ecosystems of projects or policies in a portfolio should be, at least, zero. Leaving aside, for the moment, the issue of what it is precisely that projects should (on balance) seek to conserve, the broad principles of the approaches, for example, in Barbier et al. (1990) and later in Pires (1998) for subjecting a cost-benefit test to a (strong) sustainability constraint.
A practical example of this investment constraint almost dates back the beginnings of those conceptual ideas. This starts from the premise that limits to valuation mean that certain components of natural capital (notably “biodiversity”, but not limited to this) need to incorporated within CBA as (sustainability) constraints (see, for example, Quinet et al., 2013, for a discussion of practical issues in valuing biodiversity in official CBA in France). More recent attention has focused on the form these constraints should take with, for example, an emphasis on thresholds and (safe) limits. This, in turn, requires knowledge or judgement about such thresholds across different natural assets. Another example is global climate change. Economic thinking (and CBA as part of that) has been used to help the frame discussion about what should be the appropriate level of ambition in global climate policy (see, for example, Stern, 2007; Weitzman, 2007). However, these political debates about global climate targets arguably have been based to a far greater extent on judgements about what degree of warming can be “tolerated” without physical thresholds being breached (e.g. Rockström et al., 2009; Steffen et al., 2015).
A notable practical development is the implementation of this idea in real-life settings under the guise of “biodiversity offsetting”. Common to this earlier literature, which viewed the principle of sustainability as applying to the portfolio of projects, under offsetting proposals, biodiversity – in the round – must be maintained (or enhanced) by requiring that to the extent that any one project degrades or destroys an ecosystem or damages biodiversity, this must be “covered off” by improvements or additions to ecosystems or biodiversity elsewhere: i.e. so-called shadow or compensating projects. Typically, however, the context here is one where adverse impacts are supposed to be prevented or at least minimised at the original site. It is the residual damage that is subject to offsetting via a compensating investment elsewhere (see, for example, Chevassus-au-Louis et al., 2009).
While resource compensation preserves some trade-offs between costs and benefits it plainly circumscribes cost-benefit thinking in a substantial way. Roach and Wade (2006) provide an empirical investigation of this resource compensation or “equivalency” in the context of habitats. And as mentioned, the practical counterpart is policy instruments which variously go by the names of “mitigation banking”, “habitat banking”, resource equivalency (REMEDE, 2008) or “biodiversity offsets”. A commitment to scaling up these schemes can be found variously in the Aichi targets (as part of the UN Convention on Biodiversity), EU Biodiversity Strategy to 2020 and the UK Natural Environment White Paper (Defra, 2011). BBOP (2012) defines the latter as: “… measurable conservation outcomes resulting from actions designed to compensate for significant residual adverse biodiversity impacts arising from project development after appropriate prevention and mitigation measures have been taken.” (p. 12).
Moreover, it asserts the goal of these interventions as “… to achieve no net loss and preferably a net gain … with respect to species composition, habitat structure, ecosystem function and people’s use and cultural value associated biodiversity” (BBOP, 2012, p. 13). This is a challenging array of attributes to offset. Not surprisingly, practical examples of biodiversity offset schemes have fallen short of this ambition, often relying instead on relatively simple metrics on habitat extent and quality. This has led to debates about whether compensation is genuinely “like-for-like” (and how this might be better guaranteed) as well as other issues such as governance and additionality or leakage (see, for example, Bull et al., 2013; Gardner et al., 2013; POST, 2011). A discussion of these issues, including good practice insights, is provided in OECD (2016).
What biodiversity offsetting does is to place CBA within a strategic constraint. That is, once the constraint is known then the rules for appraising the costs and benefits are relatively simple, at least in principle: CBA must work within this institution with the decision-rule being to maximise net benefits subject to observing the constraint. Needless to say, a number of questions might be posed. For example, are such constraints a special case or rather more general strategic considerations? How are these constraints to be determined?
Turning to the latter question first, given the characteristics of strong sustainability, guidance from the natural sciences, on how much of nature should be conserved, must be focal to this. Presumably social considerations play their part too, not least in crafting this technical advice to what is judged to be politically possible. How far this economic thinking, and especially judging costs and benefits, should be used for determining the strategic constraint is also important to reflect upon. Consideration of costs and benefits cannot have primacy – the point of the strong sustainability constraint is that there are clear limitations to this. However, neither can those considerations be irrelevant. Guiding principles such as the safe minimum standards potentially seek to perform this balancing act.
Another way of understanding these debates, however, is to acknowledge that these reflect different “belief systems” thought to be focal to policy problems. Put this way, CBA represents one belief system; based on an assumption of the importance of being explicit about the implications of policy choices for the way in which economic resources are used and, in particular, the trade-offs that this involves. Alternate belief systems might reject these trade-offs, perhaps by prioritising protecting nature arrived at through particular ethical perspectives. Rather than reject CBA altogether the “sustainability constraint” approach becomes a useful way of viewing the implications of these different beliefs. Not least it facilitates some explicit understanding of the costs of observing constraints (as well as the benefits).
Biodiversity offsetting is a specific constraint, albeit one which applied properly, and consistently, is possibly far-reaching. So too for carbon offsetting. But clearly, strong sustainability is a wider set of concerns about nature, and natural capital, more generally. A reasonable question is whether such constraints are special cases, or something more pervasive, and what the character of these constraints should be. Certainly the spirit of the strong sustainability approach indicates a more ubiquitous role for setting strategic constraints. Chevassus-au-Louis et al. (2009), for example, suggest a distinction between natural capital which has “intrinsic” value and natural capital which does not as a basis for this. In turn, this necessitates practical deliberations on what characteristics might allow particular resources to be categorised as one or the other.
Where constraints are recommended, what these constraints should look like is another matter. One extreme would be a set of (piecemeal) constraints for specific natural assets (e.g. biodiversity, carbon, urban air quality and so on). Another extreme would be to define natural capital in the aggregate (e.g. Helm, 2015). This would permit far more flexibility but inevitably raises questions about whether parts of the natural asset portfolio are substitutable or not. It also requires an index of natural capital, which in principle would be reflect these trade-offs.
Finally, given that strong sustainability constraints – such as those entailed in biodiversity offsetting – necessitate actual compensation, ensuring that these corresponding investments do not themselves fail arises as a subsequent challenge. While this is a question that can only really be answered through an ex post evaluation, there are a number of ex ante considerations that might help mitigate against the failure of offsetting projects. For example, CGDD (2015) identifies a number of such factors relating to technical matters of project planning and execution as well as the institutional arrangements that govern these compensating actions. This proceeds by categorising the risks of offsetting projects such as whether these restore sufficient biophysical quantities (defined against specified metrics, such that the adequacy of these chosen metrics is also relevant). Additionally, the location of what is restored – in the compensating project – relative to what is lost by the initial (natural capital depleting) project might also be relevant: that is, for example, does the former have connectivity with other (related) ecosystems?
What the preceding points suggest is that some sort of structured planning is needed in meeting this need to impose sustainability constraints on policy and project decisions. In other words, institutional arrangements are critical. CGDD (2015) note that paying attention to governance might also anticipate risks of failure of compensating projects by addressing other risks as well. This includes managing project uncertainty generally as well as scrutinising whether management plans and economic resources for long-term ecosystem management are themselves sustainable. In turn, this suggests checking that budgets and contingency funds are adequate and that subsequent management is devolved to appropriate bodies.
12.7. Concluding remarks
The notion of “sustainable development” has permeated significant parts of policy and public discourse about the environment. While there remains debate about it means for development to be sustainable, there is now a coherent body of academic work that has sought to understand what a sustainable development path might look like, how this path can be achieved and how progress towards it might be measured. While it is hardly surprising that these efforts have not generated a consensus, there has been considerable progress in understanding where agreement and disagreement is and why this arises. Perhaps the pre-eminent example of the arrival of sustainability to the policy agenda is the UN Sustainable Development Goals or SDGs.
Much of this work considers the pursuit of sustainable development to be an aggregate or macroeconomic goal. By-and-large cost benefit analysts have not sought actively to engage with this broader debate except insofar as it relates to factors affecting a project’s forecast net benefit or rate of return. However, it should be noted that recent developments discussed elsewhere in this volume – most notably on valuing environmental impacts and discounting costs and benefits – are relevant to this issue. In this chapter, we have discussed a number of additional speculations about how cost-benefit appraisals can be extended to take account of recent concerns about sustainable development.
According to one perspective there is an obvious role for appraising projects in the light of these concerns. This notion of strong sustainability starts from the assertion that certain natural assets are so important or critical (for future, and perhaps current, generations) so as to warrant protection at current or above some other target level. If individual preferences cannot be counted on to fully reflect this importance, there is a paternal role for decision-makers in providing this protection. Some have sought to characterise this according to ecological criteria while others have drawn on political precedents or it is seen as dependent on weighting decision-making heavily in favour of precaution. This raises important issues. On the one hand, there is a benefit to avoiding untoward and irreversible damage to (possibly) critical resources. On the other hand, there are opportunity costs to applying the shadow projects approach that need still to be considered.
With regards to the relevance of this approach to cost-benefit appraisals, a handful of contributions have suggested that sustainability is applicable to the management of a portfolio of projects. This has resulted in the idea of a shadow or compensating project. For example, this could be interpreted as meaning that projects that cause environmental damage are “covered off” by projects that result in environmental improvements. The overall consequence is that projects in the portfolio, on balance, maintain the environmental status quo. Practical applications of this approach arguably include biodiversity offsetting.
There are further ways of viewing the problem of sustainable development. Whether these alternatives – usually characterised under the heading “weak sustainability” – are complementary or rivals has been a subject of debate. This debate would largely dissolve if it could be determined which assets were critical. As this latter issue is itself a considerable source of uncertainty, as discussed above, the debate continues. However, the so-called “weak” approach to sustainable development is useful for a number of reasons. While it has primarily be viewed as a guide to constructing green national accounts (i.e. better measures of income, saving and wealth), the focus on assets and asset management has a counterpart in thinking about project appraisal. For example, this might emphasise the need for assess stocks of assets before the project intervention and what they are likely to be after the intervention.
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Notes
← 1. Gerlagh and van der Zwann (2002) look at the case where individuals have a very strong preference for natural assets rather than non-substitutability per se. This is very similar to the notion of a lexicographic preference that has been the subject of a mini-literature in stated preference studies. The implications of this assumption, however, are that liquidating a natural asset beyond some threshold plausibly lowers the maximum level that future well-being can take.
← 2. This approach can also accommodate a crucial concern about the nature of ecosystem assets: namely, that these resources are subject to threshold effects where services are subject to (possibly) greater risks of abrupt and extreme changes once a critical level of the asset has been breached.