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This is an extensive revision of Dan Faith's 2008 entry on Biodiversity for the Stanford Encyclopedia of Philosophy.
Faith, D.P. in review. "Biodiversity", The Stanford Encyclopedia of Philosophy, Edward N. Zalta (ed.)
“Biodiversity” often is defined as the diversity or variety of living forms, from genes and traits, to species, and through to ecosystems. "Biodiversity" was coined as a contraction of "biological diversity" in 1985, but the new term arguably has taken on a meaning and import all its own. A symposium in 1986, and the follow-up book BioDiversity (Wilson 1988), edited by biologist E. O. Wilson, heralded the popularity of this concept. Ten years later, Takacs (1996, p.39) described its ascent this way: "in 1988, biodiversity did not appear as a keyword in Biological Abstracts, and biological diversity appeared once. In 1993, biodiversity appeared seventy-two times, and biological diversity nineteen times". The 2007 revision of this SEP entry suggested that it would be hard to count how many times "biodiversity" is used every day by scientists, policy-makers, and others. In 2013, the continued increase in the use of the keyword “biodiversity” in the scientific literature corresponds to an approximate doubling of the number of biodiversity papers every 5 years. The importance of biodiversity issues regionally and globally also is reflected in the Convention on Biological Diversity’s targets for 2020 (http://www.cbd.int/sp/ ), and in the recent establishment of the Intergovernmental Platform on Biodiversity and Ecosystem Services (IPBES).
While the history of the term “biodiversity” is relatively short (compare it to other terms covered in this encyclopedia), it raises important and distinctive philosophical issues. This is reflected in interesting recent books on philosophy of biodiversity, for example by Maclaurin and Sterelny (2008) and Maier (2012). Some philosophical issues arise in the ongoing debates about the most useful definitions of "biodiversity", and in the characterisation of its intrinsic and anthropocentric values. The term “biodiversity” now is used more widely - but less clearly - than ever before. Issues include those relating to the possible relationships between biodiversity and ecosystem services (an ongoing topic for IPBES). "Ecosystem services" broadly refers to the goods and processes, and resulting potential benefits to humans, found in natural (and managed) ecosystems. Ecosystem services (for example, clean water) provide clear links to human well-being, but biodiversity itself also represents anthropocentric benefits/values. The early rationale for the term “biodiversity” (e.g. IUCN 1980, McNeely 1988) refers to the “option value” of biodiversity, reflecting the value of maintaining living variation in order to provide possible future uses and benefits.
The ecosystem services movement expands the range of possible meanings of “biodiversity”, challenging standard definitions. While any use of the term “biodiversity” logically might be expected to capture some notion of “diversity”, current usage sometimes is too specific (e.g., equating it with a single species or other element); too general (e.g., vaguely equating it with the “fabric of life” or all of nature), or too tangential (e.g., equating it with any ecological factor relevant to management for ecosystem services). Consequently, values actually tied to variation/diversity (such as the option value of biodiversity) now risk being neglected in planning and decision-making.
Ecosystem services themes focus on management and decision-making about places, but biodiversity conservation necessarily has a broader scope. It considers a wide range of possible “objects” for management and decision-making – not just places, but also species, populations, and other entities. It considers a range of possible “units” within those objects – species, genes, features, etc. Quantifying biodiversity, in principle, then amounts to counting up the number of different units within a given object, or within a given set of objects. Critically, biodiversity conservation science also recognises that individual objects have other (positive or negative) values (when the objects are places or ecosystems, these include the much-discussed “ecosystem services”). Addressing the full scope of “biodiversity” requires a workable conceptual framework, based on the core idea of “biodiversity” as the variety of life at multiple levels, and expressed in a way that promotes integration with other needs of society. This SEP entry focuses on these conceptual issues.
A sequel to BioDiversity (1988), naturally titled Biodiversity II (1997), documents the rapid rise of the term "biodiversity" in importance and influence. This book (Reaka-Kudla et al. 1997) interprets biodiversity broadly, consequently tracing some aspects of “biodiversity” back as far as Aristotle. The scientific literature over the past 20 years similarly documents broad interpretations, in which "biodiversity" sometimes is used to mean "life" or "wilderness" or even "conservation" itself. Expanded meanings may refer to a favourite single species, or to the maintenance of ecosystem processes. Within the ecosystem services movement, the term “biodiversity” sometimes refers to most any aspect of an ecosystem that helps deliver services.
These developments suggest that the term “biodiversity” forever might be pushed and pulled in different directions, depending on current fashion. However, the situation may not be as chaotic as it appears. When a formal definition is required, descriptions of “biodiversity” typically revert back to some version of that core idea of living variation, over levels from genes to ecosystems. One anchor is provided by the widely quoted definition of the Convention on Biological Diversity (CBD, article 2; http://www.cbd.int/sp/):
“the variability among living organisms from all sources including, inter alia, terrestrial, marine and other aquatic systems and the ecological complexes of which they are part; this includes diversity within species, between species and of ecosystems”.
The next sections trace developments in the characterisation of biodiversity, including: the shift from single species to holistic perspectives on biodiversity; process versus inventory interpretations; and attempts to develop a coherent framework to describe variation across multiple levels.
1.1 From single species to holistic perspectives on biodiversity
Debates about values of individual species have influenced the development of ideas about the holistic value of biodiversity (for review, see IUCN 1980; Norton 1988a,b). “Holistic” is used here to refer to any cases when we are talking about values or properties of biodiversity itself, rather than values or properties of elements of biodiversity.
In considering values of individual species, commodity value and other direct use values for species have intuitive appeal in reflecting known values. Nevertheless, species may be preserved for reasons other than their known values as resources for human use (Sober 1986). Callicott (1986) discusses philosophical arguments regarding non-utilitarian value and concludes that there is no easy argument to be made except a moral one. Species have some "intrinsic value" — reflecting the idea that a species has a value "in and for itself" (Callicott 1986, p.140, see also Soulé 1985, McShane 2007).
A philosophical issue is whether such species values depend on a human-centered perspective. Regan (1986) argues that we need "duties that are independent of out changeable needs and preferences." Callicott (1986) sees the intrinsic value of species as not independent of human values, because such values can be linked to Hume's theory of moral values (see also Norton 1986).
Randall (1988, p. 218) argues that preference is the basis for value and that it is possible to treat all species values as preferences of humans. Preferences-based approaches to valuation can provide economic estimates of value. A rationale for such approaches is that the only good way to protect species is to place an economic value on them. Randall argues that such quantification is advantageous because the species preservation option will fare well when the full range of values is included in conservation priority setting.
Many of these arguments are motivated by the desire for criteria for setting priorities among species for conservation efforts. Such priority setting requires decisions about the role of "triage" based on species prioritization. Triage recalls the medical context in which priorities are set for investments in saving patients. Applied to conservation, individual species are differentially valued and assessed relative to differential opportunity costs. The best conservation package is to be found through a process of calculating costs and benefits of protection of individual species.
Some biologists reject the idea of triage and argue that we must try to save all species (Takacs 1996). The latter option is arguably more holistic and in accord with a focus on all elements of biodiversity.
The Preservation of Species (Norton 1986) documents an attempt to move from values of species to some overall holistic value of biodiversity, rejecting typical triage arguments based on the benefits versus costs for individual species. This book could be called a "prequel" to BioDiversity (1988): the title suggests a species focus, but the book's subtitle refers to biological diversity. Norton criticizes the "benefit — cost" approaches for priority setting for species as piecemeal, because every species must exhibit actual or potential use to justify itself. He argues that every species arguably has utilitarian value and that species perceived values are hard to estimate. For this reason, trying to place dollar values is "doomed to failure" (1986, p. 202). Norton concludes that we should abandon the "divide and conquer" approach and look at total or holistic diversity, with species as the fundamental unit: "each species in an area can be viewed as a unit of total diversity." Ehrenfeld's (1988, p. 214) position is even more sharply defined: "value is an intrinsic part of biodiversity; it does not depend on the properties of the species in question."
This perspective demands some alternative to species-based triage that will still accommodate the reality of limited resources. The idea of a "safe minimum standard" (SMS) for biodiversity has been proposed as a suitable alternative to triage. Norton advocates an SMS based on unit-species, interpreted to mean that all species are saved unless costs are intolerable. Wilson (1992, p. 310) also advocates an SMS in which all species are to be protected unless costs are too high. He argues that we "treat each as an irreplaceable resource for humanity" rather than examining single species and their properties and deciding how much to invest.
The SMS leaves the idea of "too high a cost" open to different interpretations. These vary with philosophical perspectives about the nature of values. For example, "deep ecology", where biodiversity is independent of human value, responds differently to "utilitarianism", where biodiversity might be preserved to extent that measurable benefits to humans exceed costs (see The Preservation of Species). Randall's (1986, p. 103) utilitarian position considers intrinsic or option value of unit-species in conjunction with any recognized utilitarian value: all species not already distinguished in having recognised human-use values "would be treated as having a positive but unknown expected value; implicitly all would be treated as equally valuable."
Despite difficulties in actual implementation, the idea of an SMS based on species as the countable units of biodiversity remained popular from the 1986 The Preservation of Species through at least to Takacs' review (1996). Takacs review extends earlier objections to differentiating and prioritizing among species. He criticises quantitative approaches, developed in the early 90's, that make taxonomic distinctions among species. These methods reflect the idea that a taxonomically (or phylogenetically) distinctive species may deserve a higher priority for biodiversity conservation (see IUCN 1980; Faith 1992). Takacs (1996; p. 61), citing the early proposals of this kind, objects to the resulting "intricate" calculations to prioritize species based on taxonomic distinctiveness. He claims that "we can avoid tedious mathematical calculations of relative species value by switching to biodiversity" (counting up unit-species).
Takacs’ argument focuses on one level for biodiversity – species. In contrast, some early discussions of biodiversity advocate a holistic perspective on biodiversity that finds equal value units at multiple levels of variation. Callicott (1989) and others follow Aldo Leopold's (1949) work in arguing that all levels of biological organization (species, biotic communities, ecosystems) have intrinsic value. E. O. Wilson (1988) suggests values for holistic biodiversity extending beyond intrinsic values. For Wilson, biodiversity captures the idea of a "frontier of the future", presenting a dazzling prospect of largely unknown variation at multiple levels, with unanticipated uses and benefits for humans. Wilson characterises biodiversity as corresponding to a dramatic transformation from a "bits and pieces" approach to a much more holistic approach. Accompanying this is a realization that biological diversity – living variation - is disappearing and such losses are irreversible. Wrapped up in the term is the idea of a "biodiversity crisis" - a loss of living variation sometimes described as a 6th great extinction in the history of life on Earth. Ehrenfeld (1988) similarly promotes the value of diversity in the aggregate in his argument that diversity, or living variation, previously was never regarded in itself to be in danger, but now is recognised as endangered in its own right.
In accord with the idea of a "frontier of the future", biodiversity symbolises our lack of knowledge about the components of life's variation and their importance to humankind (see Takacs 1996). Biodiversity involves the two-fold challenge of unknown variety, having unknown future benefits. The term "option value" captures the idea of possible future benefits from biodiversity (for definitions, see IUCN 1980). A species, or other element of biodiversity, has option value when its continued existence retains the possibility of future uses and benefits. We also can refer to the option value of biodiversity holistically. IUCN (1980) advocated conservation of diversity in order to ensure benefits “for present and future use”, and McNeely (1988) similarly referred to “option values” of biodiversity. Reid and Miller (1989) echo these ideas in their early paper, “Keeping options alive: the scientific basis for conserving biodiversity” (see also Wilson 1992 The Diversity of Life; Faith 1992; 2012a).
The Millennium Ecosystem Assessment (MA 2005) clearly links biodiversity to option values: “Biodiversity loss is important in its own right because biodiversity has cultural values, because many people ascribe intrinsic value to biodiversity, and because it represents unexplored options for the future (option values)”. The MA describes option value as: “the value individuals place on keeping biodiversity for future generations”. Option value reflects not only to the unknown future benefits from known elements of biodiversity, but also to the unknown benefits from unknown elements. Biodiversity option value therefore links "variation" and "value": the value of maintaining variation in order to maintain possible future benefits.
The range of values associated with biodiversity can be categorised in other ways. Ehrlich and Wilson (1991) propose three basic reasons why we should care about biodiversity. The first is most closely linked to intrinsic value: “moral responsibility to protect what are our only known living companions in the universe.” Their second reason is that humanity has already obtained foods, medicines, and industrial products and other benefits from biodiversity, and has the potential for many more. Thus, this links to option value of biodiversity. Their third reason is based on the recognised ecosystem services provided by natural ecosystems. Here, they make a link to biodiversity in arguing that “diverse species are the key working parts” within such ecosystems.
The section below, on possible general frameworks for characterising biodiversity, finds some immediate links to option and intrinsic values. Later sections additionally consider the possible ways that biodiversity links to ecosystem services values.
1.1.1 "Phylogenetic diversity" (PD) as an example of units other than species
The idea that option values exist for units at multiple levels of variation suggests re-examination of the early criticisms about making taxonomic distinctions among species. Takacs’ argument is that we do not know enough about species to assign different values. Such concerns seem to be well-targeted at the early taxonomic distinctiveness methods. These typically assign differential values to species through calculated phylogenetic distances, or differences, among species. However, one of the early methods (Faith 1992, 1994) offers a different perspective on this problem. Rather than focussing primarily on differences between species, Faith interprets phylogeny as a way to talk about biodiversity with reference to some lower level units. He suggests that a phylogenetic tree provides a pattern and evolutionary model that allows inference of the amount of “feature diversity” represented by any given subset of species (a feature might be some morphological characteristic of a species). Here, “biodiversity” once again is based on equal-value units – but now the units are features of species, not the species themselves. The rationale for option value at the level of unit species extends naturally to unit features and their unknown future benefits. It follows from this argument that maintaining more features maintains more option value.
The use of phylogeny to infer feature diversity is an attempt to overcome our lack of knowledge about specific features of species. Not only do we not know, in general, the future value/benefit of different features, but also we cannot even list all of the features for most species. Phylogenetic pattern provides one way to estimate and quantify variation at the feature level. The estimated relative feature diversity of a set of species is its "phylogenetic diversity" (PD; Faith 1992, 1994). The PD of a set of species is calculated as the minimum total length of all the phylogenetic branches required to connect all those species on the tree. This definition follows from an evolutionary model in which branch lengths reflect evolutionary changes and shared ancestry accounts for shared features. The model implies that PD in effect counts-up the features represented by a set of species. Any set of species that has greater PD will be expected to have greater feature diversity.
Faith (1992) concludes that PD quantifies biodiversity option value at the level of features of species. Larsen et al. (2012) argue that “it is difficult to provide a robust proxy for ‘option value’ – the potential value to society – as these values are not yet realized”. However, they conclude that “a compelling argument can be made that maximizing the retention of phylogenetic diversity (PD) should also maximize option value, as well as diversification and adaptation of the species in a future of climatic change”.
Forest et al. (2007) explore PD and option value in their examination of the phylogenetic distribution of angiosperm plants having known human uses (classified as medicinal, food, and all other uses). They studied an estimated phylogenetic tree for nearly 900 genera found in the Cape hotspot of South Africa. Forest et al. (2007) concluded that, if we did not know about those medicinal, food, and other uses, then preserving sets of species with high PD would be a good way to preserve these unknown benefits. This argument supports the link between PD and biodiversity option value at the level of features.
Similarly, Huang et al. (2012) advocate the use of PD in conservation because they find that it provides a much stronger link to “trait diversity”, relative to species. PD now is regarded “as a leading measure in quantifying the biodiversity of a collection of species” (Bordewich and Semple 2012). The description of PD as “a resonant symbol of the current biodiversity crisis” (Davies and Buckley 2011) also highlights recognition that loss of variation is critical for units other than species.
These properties may de-fuse some of Takacs’ original criticisms. PD is not “intricate” arbitrary calculations, and is not a fruitless distraction from "biodiversity". It is all about biodiversity – but it is focussed on another level of variation. The PD framework can be reconciled, at least partly, with Takacs’ preference for equal-value units over differential values among units. The same argument used to justify species as equal-value units can be used to justify differential valuation of species (Faith 1994). Equal value for features as lower-level units can imply differential values among the species – for example, one species may offer more unique features relative to other species (Faith 1992, 1994). Thus, a paradox posed by defining equal units, at multiple levels of variation, is that equality at one level can imply distinctions among the units at another level.
PD provides a useful case study to explore the issues arising when we consider alternative definitions of biodiversity. PD is critiqued in recent studies on the philosophy of biodiversity. In one recent book “What's So Good About Biodiversity?”, Maier (2012) concludes that PD involves “demoting species to a secondary role”, and PD reflects a rejection of “a multi-dimensional conception of biodiversity”. However, the early work on PD (Faith 1992, 1994) emphasises that features of species represent just one part of a whole hierarchy of variation (contra Colyvan et al 2009). In the sections below, PD is used to illustrate some of the issues arising in the quantification of unknown variation and option value at multiple levels of biodiversity.
1.1.2 General frameworks for units at multiple levels of variation
Sarkar and Margules (2002) suggest that, when we speak of genes, species, and ecosystems, these are not the specific entities of interest; instead, they are benchmarks for the full hierarchy of variation: "there is heterogeneity at every level" (2002, p. 301). However, Sarkar (2005) argues that any consideration of biodiversity across all biological levels, from genes to ecosystems, amounts to considering all biological entities, so that biodiversity absurdly "becomes all of biology" (see also Norton 1987).
Sarkar (2005) proposes, as one practical solution, that “biodiversity” operationally amounts to whatever is the valued target of conservation priority setting for different localities (for related discussion, see Meinard and Grill 2011). This operational perspective appears to be compatible with any framework that defines some in-principle countable units of biodiversity: a locality has high conservation priority if we can count-up the additional elements of biodiversity that it contributes. A focus on countable units therefore avoids the absurdity of seeing biodiversity as "all of biology" by clearly defining a specific realm within the broader biological sciences. A related argument is put forward by Sarkar and Margules (2002, p. 302): "the relative concept of biodiversity built into the definition of complementarity has the level of precision needed to undertake conservation planning." This important concept of complementarity — the counting-up of the additional elements of biodiversity contributed – is discussed in the section below.
A general framework for biodiversity might have the property that it identifies countable units at multiple levels of variation. Maier proposes a useful general framework of this kind in which there are “categories” covering all possible levels of variation, and different units are recognised within these. For example, the category may be “species diversity” and we count of number of different species as units.
Maier’s criticisms of the PD approach in this context are revealing. He argues that, because we often see convergent evolution, the PD assumption of shared-ancestry explanations for shared-features is inappropriate. Maier also argues that PD is unsatisfactory in not allowing us to list and characterise “all features at least for extant species collectively”. This perspective contrasts with Faith’s (1992) expressed intention that any framework for talking about units of variation must make inferences about those units that are not yet known to science. Further, while PD, through the shared-ancestry/shared-features model, makes inferences about relative feature diversity, the model can “fail” for any individual feature. One obvious reason is convergent evolution, which is addressed below, in considering an alternative shared-habitat/shared-features model.
This discussion highlights the idea that there are multiple versions of objects, units and pattern-process models. Faith characterises PD as just one case of a general framework for biodiversity that uses pattern-process models to link objects and known/unknown units (Faith 1994; Faith and Walker 1996). In general, the biodiversity “units” are the things we would like to count up, and the “objects” contain various units. Typically, many units remain unobserved/unknown, and the pattern-process model defines relationships among the objects that enable inference of relative numbers of units represented by different sets of objects. Thus, PD is the specific case where species are the objects, consisting of different features as units, and the pattern and model is based on phylogeny.
The objects naturally will vary in how many, and which, units they represent. Consequently, different sets of such objects vary in how well they represent the total number of different units. The biodiversity of a set of objects is not the sum of the biodiversity of those objects in the set. The gain in units from the addition of an object ot a set of objects is given by its complementarity value to the set - the number of so-far unrepresented units. These units are inferred rather than directly counted. As illustrated for PD, measured “pattern complementarity” predicts unit-level complementarity (Faith 1994).
In another example of this pattern-model framework, ecosystems or localities are the objects and species are the different units. Again, different sets of such objects vary in how well they capture the total number of different units. One possible pattern-process model linking objects to units is based on the idea of unimodal “response” or adaptation of species to environmental gradients. Unimodal response means that shared habitat explains, and predicts, shared species among different localities/objects. The pattern here is the configuration of localities along multiple environmental gradients (an ordination). The relative biodiversity of a set of localities (objects), at the level of species (units) then is inferred by applying a p-median criterion to this pattern (the “ED” method; Faith and Walker 1996).
The gain in species from the addition of a locality to a set again is given by its complementarity value to the set. Note that “complementarity” originally was defined only if one could count up species among localities. In contrast, the ED model allows the overall biodiversity gain at the species level, from protecting an additional locality, to be predicted from the pattern; the complementarity value is predicted by the degree of increased p-median representation of the pattern provided by that locality (see also Nehring and Puppe 2002, Sarkar 2012).
In another important context for this framework, the populations of a given species are the objects, and genetic variants are the units. Minh et al. (2007) use a pattern-model based on networks, rather than phylogenetic trees, in order to make inferences at the level of units equated with genetic variants among populations.
A unimodal response model and gradients pattern also is applicable to the case where objects are species and the units are features or traits. Here, shared habitat explains, and predicts, shared traits among species. These may be the convergently derived traits (referred to above) that reflect functional diversity among species. Biodiversity at the level of functional trait diversity may be applied within an ecosystem. There are many measures of functional trait diversity (Pla et al. 2011), but many simply are based on specific nominated traits. Others use some measure of differences among species. In contrast, a measure based on shared habitat explaining shared traits (Faith 1996) nicely parallels PD’s assumption that shared ancestry explains shared features/traits (Faith 1992). This illustrates how the same objects and units may be linked by alternative pattern-process models.
In all of these examples, greater biodiversity (number of units) provides greater option value, and also may correspond to greater current benefits or services. However, because the objects overlap in their units, the same number of objects (as a subset) may represent a large or a small amount of biodiversity.
1.1.3 Difficulties in applying objects and differences approaches
The section above refers to some early taxonomic distinctiveness methods that assign differential values to species through calculated phylogenetic distances, or differences, among species. Weitzman (1992) presents a general framework for biodiversity based on this idea of objects and measures of difference between pairs of objects. The biodiversity of a given set of objects then is reflected, not in a list of the different objects, but in the amount of difference represented by the set. This approach assumes that we know enough to define meaningful differences among the initial objects. Weikard (2002), following Weitzman’s object-differences framework, argues that, “an operational concept of diversity must rely on some measure of dissimilarity between appropriately defined objects.” Maclaurin and Sterelny (2008), in their book, “What is biodiversity?”, and Morgan (2010) also see this approach as a core framework for characterising biodiversity.
One challenge for the objects-differences approach is that, even if a good definition of pair-wise difference is found, it is hard to covert this to an evaluation of a given set of objects. These issues are apparent in Weitzman’s (1992) exploration of the objects-differences framework for phylogeny or taxonomy. Weitzman’s approach here contrasts with the idea of equal value units within objects-as-species (the basis for PD). Faith (1994) compares the dissimilarity approach to PD and argues that Weitzman’s approach cannot in general evaluate a evaluation of a given set of species. Weitzman (1998) presents a modified biodiversity model, creating an equivalent to the PD measure of Faith (1992).
Recent debates about the properties of PD as a biodiversity measure illustrate other challenges in applying the objects and differences framework. MacClaurin and Sterelny embrace the idea of biodiversity option value (noting examples where apparently unremarkable species are later found to be valuable), but they find difficulties in linking this idea to the objects-and differences approach to biodiversity. Maclaurin and Sterelny interpret PD as an application of the objects and differences framework, with species as objects and differences given by “genealogical depth”. This is an incorrect interpretation. Under the PD model, phylogenetic differences between species, if calculated, would correspond to path-length distances. In any case, the PD of a set of species is not a summation of such differences.
Maclaurin and Sterelny appear to ignore the underlying evolutionary model for PD linking shared ancestry to shared features. Consequently, they characterise PD as “theoretically unmotivated”. This oversight also seems to be the basis also for their complaint that comprehensive knowledge about characters is required, but is unavailable. Maclaurin and Sterelny have particular interest in knowledge about those characters judged to be of value. They conclude that “the Achilles heel of the phylogenetic approach is the assumption that all character changes are equally important.” (Maclaurin and Sterelny 2008, p148). In contrast, for Faith (1992), this equality of features-as-units is exactly the requirement for talking about option value of biodiversity at the level of features.
Maclaurin and Sterelny go on to argue that phenetic (overall similarity) approaches that define a “morphospace” make fewer assumptions and so suffer from fewer problems than phylogenetic approaches (see also Morgan 2010). Justus (2010) notes that this kind of approach is largely constrained to inferences only about the initial characters used to define the morphospace: “MacClaurin and Sterelny emphasize, however, that morphospace distance only characterizes disparity in terms of the phenotypic properties defining the space. Disparity is therefore only tractable with respect to specific, empirically well-motivated morphospaces, which are rarely available.” This limitation parallels the limitations in using selected functional traits to make a traits space, as compared to using models to make more general inferences about functional traits (see earlier section).
These discussions also highlight concerns that the many equally plausible ways to calculate distances or differences leave us with a profusion of plausible indices. A recent review of phylogenetic diversity and conservation (Winter et al. 2012) laments that there is little basis for distinguishing among the large number of existing phylogenetic indices. Here, Winter et al interpret “phylogenetic diversity” as derived from any between-species biodiversity distance, based on phylogeny. Thus, the review highlights well the problems in choosing among different notions of differences. At the same time, Winter et al. fail to recognize PD as a model-based measure of feature diversity and associated option values, and so do not consider at all the possible advantages of an alternative pattern-model units approach.
PD also sheds light on another difficulty in applying an objects-and differences approach to characterising biodiversity. Morgan (2010) argues that, even if one has objects and some agreed natural measure of differences, a remaining problem is how to trade off more objects for less differences (or vice versa) in order to make comparisons among different conservation outcomes. Are 4 somewhat different objects better than 3 very different objects? The pattern-model framework properly side-steps this dilemma, because it does not entail an “objects and differences” trade-off. The comparisons among different outcomes are based on the inferred relative number of lower-level units, not on the number of objects, nor any sum of differences. Using PD as an example, a set of 4 species logically may have lower PD than a set of 3 species. Similarly, when objects are localities and units are species, the relative biodiversity at the species level, for a given set of localities, is inferred using a p-median criterion (the “ED” method; Faith and Walker, 1996). Compositional (species) differences among the localities could be calculated, but these are not directly useful for counting up species.
The focus on counts of units, rather than arbitrary trade-offs involving differences, makes clearer the links between biodiversity and its option values. These counts of units also might
be interpreted as quantifying intrinsic value: a conservation outcome that has preserved a higher count of equal-value units has done a better job in retaining biodiversity of intrinsic value. This accords with Callicott’s (1989) argument that all levels of biological organization (species, biotic communities, ecosystems) have intrinsic value (see also McShane 2007, Sarkar and Frank 2012).
An interpretation of units as having equal intrinsic value does raise an “intrinsic values paradox”. Species may be regarded as having equal intrinsic value. Yet PD, in regarding features as intrinsically valued, seems to assign those species differential intrinsic value (some have more features than others). This paradox recalls Soulé’s (1985) statement that “species have value in themselves, a value neither conferred nor revocable, but springing from a species’ long evolutionary heritage and potential”. Does the species having longer evolutionary heritage have more value of this kind?
If counts of the relative number of units, for any level of biodiversity, quantify both option value and intrinsic value, then biocentric and anthropocentric biodiversity values are more clearly linked. The “convergence hypothesis” (Norton, 1984; for discussion, see Minteer et al. 2011) states that environmental policies focused on human needs will converge on those policies that focus on intrinsic values of nature. The condition for this common ground is what Norton terms a “weak anthropocentrism” (or “intergenerational anthropocentrism”) that takes into account the needs of future generations. Biodiversity option value gives meaning to Norton’s hypothesis that convergence is promoted by intergenerational anthropocentrism. Policies focused on maintaining biodiversity option value may converge on those focussed on intrinsic values of biodiversity because these two values are expressed using the same equal-value units. This convergence issue is discussed further below with reference to ecosystem services.
1.2 Biodiversity calculus and conservation planning
The counting up of units for any measure of biodiversity within the pattern/process models framework allows for various calculations useful for decision-makers. For example, gains and losses of biodiversity at the unit level can be calulated. Many of the common biodiversity calculations in ecology at the species-level can be transformed into corresponding meaures based on other units.
Sarkar (2008) traces links from traditional species-level ecological diversity measures to measures of biodiversity (see also Magurran and McGill 2011). These measures typically are referred to as “diversity” or biodiversity measures, adding to the complexity of possible meanings of the term “biodiversity”. An alternative way to interpret this rich variety of possible calculations is to focus on a given units-based measure of biodiversity, and then refer to the many other possible calculations that can be based on those counts. As examples, complementarity, endemism, and dissimilarities between objects all can be calculated, but are not strictly measures of “biodiversity”. In principle, every index conventionally defined in ecology at the species level has its counterpart for other biodiversity units. Counting-up the relative number of units remains the core measure of “biodiversity” but the other calculations capture other aspects – for example, expected change in biodiversity as a result of extinction.
A broad family of PD calculations is derived from the interpretation of PD as counting-up features. These extend conventional species-level indices such as complementarity to the features level. A species complements others in representing additional evolutionary history (Faith 1994), as depicted in the branches of the estimated phylogeny. The degree of complementarity reflects the relative number of additional features contributed by that species. For example, given some subset of species that are well-protected, and two species in that taxonomic group that are endangered, the priority for conservation investment may depend on the relative gains in feature diversity (the complementarity values) expected for each species.
Useful PD calculations for geographic localities include PD-endemism and PD-dissimilarities between places. Also, “expected PD” calculations are based on either probabilities of extinction or presence-absence. For example, species' estimated extinction probabilities indicate amounts of “expected PD loss” (for discussion and examples, see Faith 2008).
Calculations such as complementarity are critical to biodiversity assessments and conservation planning. Sarkar and Margules (2002, p. 302) argue that, if we are considering conservation actions in different places, then "the relative concept of biodiversity built into the definition of complementarity has the level of precision needed to undertake conservation planning."
Sarkar and Margules here describe biodiversity as rooted in place, but this is just one scale of decision making. The same complementarity principle applies to other objects, counting gains in the number of units when a new object joins a set of objects.
Thus, while “complementarity” conventionally refers to species (as units) distributed among places, it equally can refer to, for example, functional traits (as units) distributed among species. This unites what had long been two independent uses of the term “complementarity”. SCP provides one usage, but ecosystem workers commonly refer to functional “complementarity” among species within an ecosystem (for a recent example see MacDougall et al. 2013). When that complementarity among species is interpreted as counting functional traits, it amounts to the same kind of “complementarity” that is used in standard SCP.
A core biodiversity conservation challenge arises from the fact that individual objects typically are the focus of decision-making, but assessment of total biodiversity depends on sets of objects. We can apply complementarity to find the best object to add to an existing set. However, the choice of the best object typically is more complicated than that. We face a second fundamental biodiversity conservation challenge. It arises from the fact that the objects typically also have some specific services/benefits or “object values”. These also should have a say in the choice among objects. Planning and decision-making therefore works with objects that have their own specific values, plus some dynamic biodiversity complementarity contribution to a set of objects.
When the objects are geographic places, “ecosystem services” – the services or benefits provided by natural ecosystems – represent one specific (well-known) example of “object values”. Systematic conservation planning (SCP) addresses these challenges, with well-established existing methods for the case where the objects are localities of some kind. SCP seeks trade-offs and synergies among biodiversity, ecosystem services, and other needs of society. The framework in the simple case views ecosystem services as a co-benefit or as a cost.of conservation (Faith 1995). For a given locality, higher complementarity, combined with higher co-benefits and/or lower opportunity costs of conservation, implies greater priority for conservation.
SCP in principle can recognise a range of possible states or conditions or land-uses for localities. A given state may offer some ecosystem services or other benefits, combined with only "partial protection" of biodiversity (implying some lower complementarity value; Faith 1995; 2012b). The preferred state for a given locality then depends on other allocations of states in the region. This balancing of biodiversity, ecosystem services and other benefits, both within and among localities, produces greater net benefits for the region. (see example here).
"Ecosystem services" conventionally refer to services from ecosystems at the "natural end of the spectrum" (Daily 1997, p.2). Human-use, transformed, places sometimes provide partial protection along with other benefits/services (call them “transformed-system services”). SCP integrate can integrate biodiversity conservation with ecosystem services and transformed-system services (offering “partial protection” of biodiversity).
The Millennium Ecosystem Assessment (MA 2005, p.143) highlights this approach in the context of biodiversity policy options:
"...an integrated biodiversity trade-offs framework (Faith et al. 2001a, 2001b) suggests how such partial protection (for example, from private land) can contribute to the region’s trade-offs and net benefits." However, the MA (2005, p.122) also observes that "The great uncertainty is about what components of biodiversity persist under different management regimes, limiting the current effectiveness of this approach." This need for information about the biodiversity contributions from natural and transformed lands complements the well-recognised focus on determining ecosystem services benefits from relatively intact places. Thus, transformed-system biodiversity contributions deserve as much attention as the popular ecosystem services, in carrying out balanced planning.
Balanced planning using SCP in principle achieves biodiversity conservation with greater co-benefits and lower conservation opportunity costs (in terms of other needs of society). Consequently, take-up of this efficiency offered by SCP can mean a reduced rate of biodiversity loss in a region (Faith and Ferrier 2005). Simply put, land-use planning and other decision making that more efficiently balances conservation with other needs of society implies reduced biodiversity losses, compared to business-as-usual (see example figure here).
These basic SCP concepts extend to other objects/units, providing a generalised SCP. Here, SCP allocates the objects, within some planning universe, to different states/conditions. It balances the need to maximise total biodiversity (count of units among objects) and the need to maximise other benefits that also are available from the objects. Sometimes, the state/condition is simply a choice between conservation or conversion to some non-conservation use. On other occasions, partial protection options exist, where the state of an object amounts to a compromise within that object between number of units retained and amount of other benefits delivered. Using PD again as an example, SCP in this case balances conservation of features-as-units with the delivery of other benefits from the species-as-objects (preserving both current and possible future evolutionary or evo-system services; Faith et al. 2010).
Ecosystem services are relevant to several levels of generalised SCP. For conventional SCP, the objects are places, units are species, specific object values include ecosystem services, and the planning universe is a geographic region. The different possible states of the objects/places indicate which units are retained and which services/benefits are delivered. The region achieves, through SCP, some total biodiversity (and associated option value) along with other benefits. In contrast, we can consider SCP where the planning universe is within an ecosystem/locality. The objects might be species and the units functional traits. The management planning for the ecosystem then tries to achieve high total functional traits biodiversity (and associated option value) along with other benefits, including delivery of functions and services. As noted above, this context employs complementarity at the level of functional traits among species.
A focus on trade-offs and synergies, across all levels of biodiversity, is the foundation for a general sustainability analysis framework (see for example, Biodiversity and regional sustainability analysis ). This framework shifts the conventional focus on conservation targets defined for individual factors to a focus on goals directly linked to achieving (or at least maintaining the capacity for) balance and net benefits. Faith (1995, p.9) “introduces an alternative to the simple targets or ideals based on individual standards for different criteria, replacing these with targets defined by optimal trade-off solutions. Such a trade-off target means that the actual value on any one criterion may well be lower (better) than that found in the best overall trade-off solution.” An improvement in one factor may not be compatible with achieving the best balanced outcome. This kind of strategy therefore counters conventional “inclusive wealth” and similar scoring approaches that simply add together scores for individual factors (Faith 1995).
The next sections consider some current challenges faced in integrating biodiversity and its option values into such strategies.
1.3 Challenges for a holistic biodiversity concept
1.3.1 Narrow definitions of option value
Maier (2012), in his book, “What is so good about biodiversity”, criticises recent advocacy of biodiversity option value, focussing particularly on MacClaurin and Sterelny’s (2008) arguments for its importance. MacClaurin and Sterelny use the term “option value “ in much the same sense as found in earlier work (e.g., IUCN 1980; McNeely 1988): the value of maintaining something in the absence of knowledge about its future benefits. In contrast, Maier interprets “option value” in accord with a traditional economics usage that refers to the option value associated with a given resource. In this context, any quantification requires a range of estimates and calculations – including, for example, estimates of reliability of stock, of risk aversion in the specific context, and of premiums people are willing to pay. Maier complains that these basics are missing in MacClaurin and Sterelny’s arguments.
In considering ecosystem services, Daily (1997) regards option value as hard to measure for much these same reasons. This consideration of option value in the context of ecosystem services further complicates definitions. Ecosystem services valuations may include option value, but typically these relate to possible future use of specific known services (e.g. future timber from a forest area). For example, DIVERSITAS (http://www.diversitas-international.org/resources/glossary ) links option value to the "availability of a particular service for use in the future". Norgaard (2010) makes similar observations about known ecosystem services as compared to the possible services possibly needed by future generations.
The stark contrast between MacClaurin and Sterelny’s and Maier’s interpretations of option value suggests that distinguishing between the specific resource/service versus “biodiversity” versions of the term might overcome the failure of existing terminology to adequately capture the different meanings. Reference to “option value of biodiversity”, for example, in the context of PD, then does not have to be interpreted to mean that the actual “value” of the options is determined in any absolute sense. Instead, over a range of conservation outcomes, there are more or less options represented, as indicated by implicitly comparing the counts of units. A set of objects representing lots of different units has lots of biodiversity and therefore lots of option value. A set of objects with less biodiversity offers less option value. This perspective seems to be consistent with the references to “option value” of biodiversity in the MA and elsewhere (referred to above).
The idea that biodiversity conservation scenarios correspond to greater or lesser option value may help to counter unrealistic interpretations of biodiversity option value as an all-or-nothing property. For example, Mace et al. (2010, p. 3) claim that a goal ‘to maintain biodiversity so as not to foreclose any options open to future generations . . . would entail a goal of no overall loss of biodiversity. . . we suggest this is unlikely to be achievable’. In reality, a zero-loss scenario is seldom considered.
1.3.2 Ignoring option value – seeing biodiversity as all about intrinsic value
Arguments for conservation of ecosystem services sometimes adopt a caricature of biodiversity that ignores biodiversity option values. For example, Balvanera et al. (2001, p. 2047) acknowledge a simple caricature pitting biodiversity for intrinsic value against ecosystem services for human well-being. Option values of biodiversity are ignored in this simplistic dichotomy.
Egoh et al. (2007, p.719) similarly state that “biodiversity and ecosystem services are associated with different values (intrinsic vs. utilitarian).” Haines-Young and Potschin (2010, p.132) argue: “Biodiversity has intrinsic value and should be conserved in its own right. However, the utilitarian arguments which can be made around the concept of ecosystem services and human well-being are likely to become an increasingly central focus of future debates about the need to preserve ‘natural capital’.” Hardy (2008, pp. 3,4) similarly concludes that “The idea of ecosystem services allows for acknowledging more than the “intrinsic” value of biodiversity by expanding the breadth of the conservation argument to include the “utilitarian” values of nature.” Thus, the argument is that only through ecosystem services do we move beyond the intrinsic values associated with biodiversity and consider utility and human uses (see also Daily 1997; Polasky et al. 2012). Mooney (2010, p.37) argues that the ecosystem services approach provides a new way to conserve biodiversity, proposing that “we need to use a better framework that can give the policy community some clear guidelines and a rationale for protecting biodiversity”.
Sometimes these arguments are extended to suggest that the debate about values is responsible for a harmful deep divide in conservation, and that this can be bridged by adopting ecosystem services conservation as an umbrella activity that sweeps along many of the intrinsic values of biodiversity. For example, Polasky et al (2012, p.140) argue that there is a “deep divide” in conservation corresponding to an intrinsic-value versus human well-being focus (see also Reyers et al. 2012a,b). They report (p. 157) that “in general, investing in conservation that increases the value of ecosystem services is also beneficial for biodiversity conservation and vice versa”. They argue that such convergence is not surprising, “given the importance of biodiversity to maintaining ecosystem services”. Polasky et al. record only within-place biodiversity, and add-up these values over multiple places. Consequently, the study has only a weak link to regional biodiversity (over sets of objects/places) and associated option values. The case for ecosystem services as an umbrella for biodiversity again is weak because biodiversity option values are not considered.
Mace et al. (2012, p.19) similarly argue that “the relationship between biodiversity and the rapidly expanding research and policy field of ecosystem services is confused and is damaging efforts to create coherent policy.” Mace et al (2012, p.20) refers to “…an intrinsic value for biodiversity, whereby organisms have value that is by definition unquantifiable and therefore nontransactable”. They also argue that intrinsic values typically are linked to iconic species, and these can be captured through corresponding ecosystem services. Mace et al conclude (p. 24):
“the fact that traditional species-based conservation can be embodied within an ecosystem services framework, shows that an ecosystems approach is not an alternative paradigm to current conservation but simply that current conservation is part of a bigger picture.”
This conclusion seems to rely both on interpreting intrinsic value narrowly and on ignoring option value of biodiversity. An alternative perspective (see earlier section) includes option and intrinsic values, and sees biodiversity conservation and ecosystem services as part of “a bigger picture” provided by SCP type approaches. In support of this perspective, Faith (2012a,b) reviews the recent tendency to ignore biodiversity’s option values and characterise biodiversity as primarily about intrinsic, non-anthropocentric, values. He argues that biodiversity option values form another important element of human well-being and these can be integrated with ecosystem services, using SCP.
1.3.3 Biodiversity–human-well-being links as a “new” strategy
The cases described above illustrate common arguments based on a false dichotomy which pits biodiversity for intrinsic value against conservation for human wellbeing. They also highlight a real distinction that can be made among all the recent discussions – some recognise, and some ignore, biodiversity option value as a basis for human well-being (see also Faith 2012a,b). The cases noted above also suggest that those who do not acknowledge option values may then claim that a human well-being focus for conservation is new and innovative.
Caricatures of biodiversity sometimes portray the links between biodiversity and human well-being as a new product, emerging from the ecosystem services movement. For example, Kareiva and Marvier (2012, p.962) characterise conservation biology as in need of a new, human well-being, orientation: “Soulé’s original delineation of conservation biology is in need of a broader framework that we label conservation science to distinguish it from an enterprise concerned solely with the welfare of nonhuman nature." Others would recognise the early development of “conservation biology” as already providing exactly that “broader framework”.
An earlier section refers to the standard integration of biodiversity and option value into a systematic conservation planning (SCP) framework that potentially considers many different aspects of human well-being. Sarkar (2012) considers SCP in his review of the history of Conservation Biology, documenting how this young discipline (following Soulé 1985) rapidly developed multidisciplinary perspectives. Sarkar describes how Conservation Biology pioneered trade-offs analyses that consider biodiversity and human well-being (e.g. as reviewed in Orians and Soulé’s edited 2001 book, “Research Priorities for Conservation Biology”). Kareiva and Marvier promote the idea of an existing divide in conservation, but that particular divide does not appear to exist.
Another supposed new perspective is found in the influential "Love not Loss," study by Futerra (www.futerra.co.uk/downloads/Branding_Biodiversity.pdf ). Their philosophy that (p. 20) “you can help the biodiversity that you love” encourages advocacy regarding the aspects of nature that are providing you with goods and services that you love. The Futerra approach is titled “the new nature message” (p. 1). However, if there is a “new” aspect here, it does not appear to be the promotion of love-of-nature (an established strategy). Instead, the actual new aspect here appears to be an intended elimination of messages about intrinsic values of biodiversity and biodiversity loss. Futerra argues that (p. 13) “intrinsic value arguments imply an obligation, one which most people have found it easy to ignore. Simply amplifying that obligation message with further evidence of mass, irrecoverable species loss isn’t likely to change anything”. They also criticize (p. 6) “the ‘biocentrics’ who believe nature has value beyond, or equal to, human value....The biocentrics are a small minority of the population, but seem to control a large proportion of biodiversity communications.” This philosophical perspective clearly dictates what not to do, including avoiding any mention of biodiversity loss and biodiversity intrinsic values: “Make sure you don’t slip back into selling conservation for nature’s sake” (p. 23).
The Futerra philosophy is troubling also because its implicit neglect of biodiversity option values suggests that the “biodiversity that you love” will not include needs of future generations. It does not appear to include any message along the lines: “you should help the biodiversity that your children’s children will love.”
The “love not loss” philosophy seems to sit awkwardly between a rationale for education and a rationale for conservation action. The love message surely is a well-established way to advertise biodiversity - but how valid is the exclusion of loss and intrinsic values messages when promoting conservation action? Within an action and policy perspective, IUCN seems to adopt a more flexible version of the “love not loss” program. It appears to include the consideration of intrinsic and option values, in order to link this program to the CBD Aichi Target 1 – that “people are aware of the values of biodiversity” (http://www.cbd.int/sp/ ).
Vermeulen and Koziell (2002) advocate an approach similar to that of Futerra in that they seek to discard biocentric/intrinsic and global option values of biodiversity. They see such biodiversity values as ignoring important local values of biodiversity, relating to ecosystem services. They argue that treating biodiversity as one composite property corresponding to global values is not helpful, and is a consequence of the fact that "the global consensus is that of wealthy countries" (2002, p. 89). Vermeulen and Koziell recommend the consideration of biodiversity in terms of services derived from it, and not as an end in itself. They claim that "the most useful biodiversity assessments are those based locally" (2002, p. 83).
A local, within-place, focus for conservation traces back to early ecosystem services papers. Ehrlich and Mooney (1983) referred to ecosystem services and “biotic diversity” within an ecosystem. Ehrlich and Wilson (1991) propose three basic reasons why we should care about biodiversity. The first was ethical and esthetic, linked to “moral responsibility to protect what are our only known living companions in the universe.” Their second reason was that humanity has already obtained foods, medicines, and industrial products and other benefits from biodiversity, and has the potential for many more. Thus, this links to option values of biodiversity. Their third reason was based on the recognised ecosystem services provided by natural ecosystems. Here, they link biodiversity to processes in arguing that “diverse species are the key working parts” within such ecosystems.
Sterling et al. (2010, pp. 1094-1095) take the notion of “key working parts” further. They suggest that we need a shift to a process-based definition of biodiversity, because a traditional focus on attributes and patterns ignores humans: “Historically, views and measurements of biodiversity often focused on characterizing the attributes of observable patterns… and afforded less attention to processes…. Often, these views of biodiversity see humans as separate from the rest of nature…”
The next section explores the roots of the within-ecosystem focus, where “biodiversity” is equated with processes and/or the ecological factors that underpin services.
1.3.4 Interpreting biodiversity as process
Norton (1994) argues that there will never be a single "objective scientific definition" of biodiversity, in the sense of a prescription for how to measure it. Norton claims that any increase in our understanding of biodiversity will make it less likely that there will be a single objective measure. This biodiversity pluralism is based on an argument that inevitably there are many different "theory bound" versions of biodiversity and many different ways to value it. This perspective is in accord with recognition of functional versus compositional perspectives on biodiversity. For example, Norton (1994, 2001) points to recent emphasis on processes that under-pin ecological "health" or "integrity"; these are seen as going beyond a conventional elements-oriented perspective for biodiversity.
Callicott et al. (1999) similarly review the standing of "biodiversity" as a normative concept that remains ill-defined. They suggest distinctions between "functional" and "compositional" perspectives. "Functional" refers to a primarily concern with ecosystem and evolutionary processes, while "compositional" sees organisms as aggregated into populations, species, higher taxa, communities, and other categories. Valuation is to encompass all of biodiversity but through a functional perspective, shifting the focus to ecosystems processes (see also Norton 1994, 2001).
Norton (2001) summarizes the development of the process perspective on biodiversity by describing three phases of growth in "biological resources" conservation over the past years. The first was the focus on individual species. The second phase was a "problematic" perception of biodiversity as all about protection of "objects" — merely expanding the list of "items" from the first phase. Here, Norton (2001, p.77) objects to an "atomistic" bias of western culture towards objects. He argues that biodiversity has been wrongly focussed on "inventory" of species, genes, ecosystems and has neglected processes that create and maintain natural values. This inventory perspective is described as "static", not dynamic (see also Frankel and Soule 1981; Takacs 1996).
Norton argues that the inadequacy of this second phase, being "ill-suited" to an emerging process orientation, has lead to the third phase based on ecosystem processes. Here, values are not to be attached to objects; instead, we should value (or "abhor") processes. The focus is on maintaining functions of healthy ecosystems, such as provision of clean air and water. Norton (2001, p.88) sees the process focus as replacing the "increasingly obsolete" inventory/items perspective of biodiversity, arguing that we "will likely move away from the inventory-of-objects approach altogether".
A focus on processes naturally influences the way in which “biodiversity” policy is framed. Norton concludes that: "…applied to biodiversity policy, we can focus on the processes that have created and sustained the species and elements that currently exist, rather than on the species and elements themselves" (2001; p. 90). More recently, Mooney (in a 2012 interview), argues that “We need to think of biodiversity in terms of how species interact together, to capture water, light, nutrients… and how in doing this they provide services to society” (http://unu.edu/publications/articles/the-true-value-of-ecosystem-services.html).
A similar policy perspective is seen in the proposal by Perrings et al. (2010) that critical or important ecosystem services can define the “biodiversity” of interest for conservation. They summarize this ecosystem services paradigm as follows (p.323): "what and how much biodiversity should be targeted for conservation depends on what services are important to maintain and with what reliability."
Related arguments focus also on ecosystem functions and processes, but use the term “biodiversity” to refer to a variety of ecological factors linked to processes and services. For example, Diaz et al. (2009, p. 55) describe “biodiversity” as "the number, abundance, composition, spatial distribution, and interactions of genotypes, populations, species, functional types and traits, and landscape units in a given system".
Faith (2011a) argues that such measures of biodiversity may simply be re-expressions of the services, in terms of their ecological basis (such as abundance and species' interactions). He argues that ecosystem services conservation then becomes an end in itself, with unknown consequences for actual biodiversity loss (the loss of units among objects). This criticism echoes early concerns that ecosystem services conservation will value and support “biospecifics” (specific elements or processes), not biodiversity (Faith 1997).
1.3.5 The fate of biodiversity units within ecosystems
The sections above described the challenges posed when biodiversity is portrayed as intrinsic-only, or is thought to be swept along by ecosystem services conservation, or is defined-away as processes and ecological factors that underpin ecosystem services. However, other challenges arise even when biodiversity units are recognised and conserved, within a place, as part of ecosystem services conservation. The key issue is that the local conservation may not serve the needs of global/regional conservation.
As noted by the MA (2005) and others, management for some ecosystem services within an ecosystem, or other locality, may imply loss of biodiversity. However, provision of other ecosystem services may depend on the overall biodiversity within the given ecosystem or locality (Cardinale et al. 2012, Hooper et al. 2012). This link to biodiversity is surprisingly hard to pin down. A review by Cardinale (2012) notes that the claims appear to outweigh the available data as to whether biodiversity – in the sense of variation - is required for the provision of ecosystem goods and services. Part of the information problem is that the many studies that address this issue vary widely in their definition of the term biodiversity. Many studies include aspects of ecology, such as abundance and interactions of species, in their definitions of biodiversity.
Enlightened management for ecosystem services may recognise that the maintenance of
option values provided by local biodiversity maintains provision of processes and services. For example, Turner (1999) argues that the number of species is a valuable index of ecosystem “reliability”. He uses the term "insurance value" rather than "option value", to refer to a kind of insurance against the failure of ecosystems to provide goods and services (see also Smith 2010).
An important scenario is one in which ecosystem services are an added benefit available from intact land (or other objects) which retains its biodiversity. Ecosystem services of this kind represent significant added benefits gained from maintaining intact places (Daily 1997). In this context, the difference between the within-places and among-places perspectives on biodiversity conservation presents a challenge. An apparent win-win scenario is one in which management for ecosystem services within a given place maintains essentially all the biodiversity elements. However, the particular units of biodiversity contributed by that place may not be the ones most needed for biodiversity conservation at the among-places scale. In general terms, the objects or places with known benefits may or may not also contribute to biodiversity of the collection – the total number of lower-level units represented.
This discussion further highlights distinctions between option value and current benefits. Maier (2012) complains that the option value is unrealistic because it implies an expectation of only good values, yet sometimes objects have values contrary to human well-being. This difficulty is resolved if one regards biodiversity option value as only capturing the idea of maintaining variety for possible future use. This does not pretend to guarantee anything about the properties, positive or negative, of any individual element. This argument reinforces the need to have on the decision-makers table both biodiversity values and specific “object values” Thus, in decision-making, we may evaluate an object by taking into account both its biodiversity complementarity and its other recognised current values.
Returning to the case of species-as-objects, an “object values conjecture” is suggested by Forest et al.’s examination of the conservation of PD and how that relates to conservation of currently-useful taxa. They find that a set of species selected for protection based on specific services/benefits may not do well at representing the feature diversity of all species. Stated generally, the protection of lots of objects with high current “object values” may not be an adequate way to protect overall biodiversity option values.
Faith and Pollack, (in press) use PD to further explore this problem of balancing among biodiversity and “biospecifics”. They find that conservation of phylogenetically clumped current-use species can reduce the capacity to retain high PD within a given budget. These considerations pose a biodiversity conservation paradox: pointing to lots of current uses is a very good way to advertise and gain support for biodiversity (as advocated by Futerra) – but focussing conservation action on those current uses is not necessarily the pathway to conserve overall biodiversity. Paradoxically, conserving more known-use species can reduce the capacity to conserve PD and its broader biodiversity values. SCP (see section above) avoids this problem by balancing the conservation of currently-valued species and the conservation of overall phylogenetic diversity (PD). Faith and Pollack conclude that any effective “sustainable use” strategy tries to preserve not only known uses but also the sustained capacity to find other uses, in other species.
The same issue arises when ecosystems or localities are the objects, and species are the different units. A set of localities selected for protection based on specific services/benefits may not do well at representing the feature (species) diversity of the whole region. The ecosystem services of a locality may well value exactly what makes that place similar to many others, even though this amounts to ‘‘redundancy’’ at the regional scale (MA 2005). A case study (Faith 2012c) shows how an increase in ecosystem services conservation, combined with a conservation budget, can mean that the region’s capacity for achieving biodiversity conservation collapses.
A good way to summarise these issues is to reconsider the early arguments for ecosystem services. Daily (1997) explores a simple scenario in order to illustrate the importance and rationale for ecosystem services. A spaceship, with limited storage space, is to be prepared in order to set up life on the moon. Daily poses the question, “which of earth’s millions of species do you choose?” Daily begins by selecting species providing known uses and “life-support functions”. However, this spaceship ultimately might be doomed – selecting the known useful species could use up all the limited space, and so foreclose including species that provide unanticipated future uses. Daily’s list of criteria does include reference to “maintaining biodiversity”; how that is balanced with selection of known-use species must be a core issue for such a spaceship.
These issues also reveal a biodiversity conservation conundrum – ecosystem services can make conservation of a place seem better than conversion – but if there is a budget (say, the total amount of land that can be retained as intact), biodiversity conservation regionally can suffer as a consequence of that ecosystem services focussed conservation. Stated in general terms, we identify specific elements (units) with identifiable benefits, but conservation of biodiversity and its intrinsic and option values cannot proceed by only focussing on those elements. Given that objects are the focus of decisions, and assuming that intact objects (retaining at least some of their units) also have benefits to humans, then we have a tempting pathway. But the pursuit of specific benefits from the object may mean loss of its elements of biodiversity; may mean retention of only some elements that are not needed in the broader set of objects, or may mean that all elements are retained but are redundant with those already offered by other objects.
This recalls the biodiversity conservation paradox: pointing to lots of objects with current uses and benefits is a very good way to advertise and gain support for biodiversity – but focussing conservation on these current uses is not an assured pathway to conserve biodiversity.
An alternative is to pursue balanced trade-offs (and synergies) among current benefits and biodiversity option values. This often will be a local versus global story. As long as local values and opportunities, whatever their source, are given weight in these trade-offs, there is no need to try to define (or re-define) the "important" values of biodiversity as local not global. Apparent conflict is resolved also by realizing that often the local values and opportunities have little to do with the biodiversity (biotic variation per se) of the place (though they typically will link to its "biospecifics").
A trade-offs perspective based on biodiversity complementarity suggests that there is good capacity for balancing different values in setting priorities in a given region. Every place has biodiversity, but its contribution to the global option values of biodiversity is indicated by its complementarity value, not its total diversity. It is the comparison of the place's current complementarity value to the other values/opportunities in that place that matters when considering trade-offs at a regional scale. There may be apparent high conflict in a region, in that places with high biodiversity have high values for some other land use opportunity, but in such cases the region may well be able to satisfy both needs. Trade-offs applications based on complementarity have suggested that other values can be integrated without much penalty to biodiversity goals
1.4 Summary remarks on concepts
The meanings (and values) of “biodiversity” vary among many partly-overlapping interpretations of the term. Popular interpretations include biodiversity as all-of-nature, biodiversity as largely about intrinsic values, and biodiversity as a favourite species. Maintaining a suite of ecosystem services requires good ecology, and it seems to be a small, but miss-calculated, leap to say that this is the study of “biodiversity” – a consequence is that biodiversity becomes most any ecological factor supporting ecosystem services.
One of the key issues highlighted in the sections above is the question of convergence. Some claims for convergence may simply be a product of definitions. When “biodiversity” is defined as those aspects of ecology that support ecosystem services, policy for ecosystem services conservation by definition also supports biodiversity. On other occasions, a proposed strategy promoted as convergence in reality may be a thinly-veiled strategy for “substitution”. Policy focussing on what locals “love”, and not on intrinsic value and loss, might be thought to converge on overall biodiversity conservation – but, in reality, may amount to a strategy to dismiss concerns about global biodiversity option values.
As discussed above, even if local love interests perfectly preserve local elements (units) of biodiversity, this may not preserve regional biodiversity - convergence defined locally does not guarantee convergence regionally. Regional conservation planning examples suggest that, even when ecosystem services conservation ensures persistence of local elements of biodiversity, this may not ensure effective regional-scale biodiversity conservation.
The idea that policies focused on human needs and future generations converge on policies that focus on intrinsic values is plausible – under one condition. That condition is that the human-needs oriented policy includes biodiversity option value and future generations. In contrast, if human-needs policy is based only on current ecosystem services, convergence seems less plausible. It may be useful, therefore, to look beyond Norton’s “convergence hypothesis.” A complementary “sustainability convergence hypothesis” might claim that policies focused on ecosystem services and other human needs will converge on those policies that focus on biodiversity option and intrinsic values - if both aspects are “on the table” for balanced regional-scale decision-making. In this way, synergies and efficient trade-offs can be found, and the balanced policy looks a bit like both targeted policies (for example in land use planning allocations). In other words, the two policies look more alike because there is enough flexibility in the system to find solutions that accommodate both.
The requirement that biodiversity and other aspects of human well-being are “on the table” for decision-making raises challenges to fill critical knowledge gaps. The Millennium Ecosystem Assessment [Millennium Ecosystem Assesment web pages] called for further work on developing a calculus of biodiversity, so that SCP-type trade-offs approaches can better integrate a wide range of biodiversity conservation instruments (e.g., protected areas; payments to private land owners; control of invasive species). The next section addresses challenges in the growth of biodiversity-related knowledge.
The sections above highlight the role of complementarity — the additional contribution made by a place (or other object) to the overall representation of biodiversity. The true biodiversity complementarity of a place inevitably is unknown, and must be estimated using some surrogate or proxy information. For example, at a whole country scale, a simple surrogate may use available species data to estimate the biodiversity contributions of different places. Sometimes a whole country study may not focus directly on variation at the genetic or even species scales, but might use ecosystem types or similar as the surrogates to assess the representativeness of its protected areas system.
Sarkar and Margules (2002) discuss the role of biodiversity surrogates, arguing that even the relative concept of complementarity has a "conventional" element built into it because it relies on "estimator" surrogates (say, a set of butterfly species) for "true" surrogates (say, the use of species as the basis for assessing biodiversity complementarity of places). They argue that estimator surrogates are subject to empirical justification, but true surrogates remain dependent on convention. They defend this conventional element: "a philosophical point, widely appreciated by philosophers of science, but often not explicitly acknowledged by scientists, deserves to be noted in relation to this: conventional elements almost always enter into theoretical reasoning in science (Nagel 1961, Sarkar 1998). But "conventional" does not mean "arbitrary": it means that there were choices to be made, no single option was dictated by the facts at hand, and a choice was justified instrumentally by its ability to achieve the purpose for which it was intended" (2002, p. 307).
Development of empirical approaches for determining effective "estimator" surrogates for biodiversity raises philosophical issues as well. There are plenty of observations about the congruence (or not) of surrogates with other components of biodiversity (for review, see Faith 2003), but what constitutes good evidence for an effective biodiversity surrogate?
Popperian corroboration provides one pathway to assess evidence for such hypotheses. Corroboration is attractive because it does not attempt to assign probabilities of truth to hypotheses (Popper 1982, p. 346). Instead, it focuses on the evaluation of the particular evidence at hand. Proposals have focussed on the idea that corroboration assessment asks whether apparent good evidence for an hypothesis is "improbable" without the hypothesis — it cannot easily be explained away by other explanations (possible explanations suggested by our "background knowledge"). Popperian background knowledge is assigned an important role in this interpretation — the investigator is obligated to try to discover any background knowledge that would suggest that the evidence is probable even without the hypothesis (for discussion, see Faith and Trueman 2001; Faith 2006).
The interest in the role of corroboration in biodiversity studies prompts debates about its role and meaning. As background to these issues, it is revealing to examine one of Popper's own examples of falsification/corroboration, as presented in the entry on Karl Popper. This example, based on the discovery of the planet Neptune, effectively highlights the limited prospects for actual falsification, but may be under-appreciated as an example of corroboration. The hypothesis of interest in this example is Newton's theory, and the evidence is the observation of the new planet, in a position predicted by this hypothesis. Popper (1982, p. 247) argues that "a moving star, planet, would have been significant, because unexpected." Popper argues, "the unexpectedness of an event can be identified with a low probability, in the sense of the calculus of probability, on the background knowledge" and that the "predictions which lead to the discovery of Neptune, were such a wonderful corroboration of Newton's theory because of the exceeding improbability that an as yet unobserved planet would, by sheer accident, be found in that small region of the sky where their calculations had placed it". Corroboration was achieved because "the success of the prediction could hardly be due to coincidence or chance".
This example supports the idea that Popperian corroboration for a biodiversity hypothesis arises only if the evidence is judged to be improbable — in spite of attempts to identify background knowledge that suggests that the evidence is probable even without the hypothesis.
For biodiversity surrogates, a common hypothesis is that the pattern of species "turnover" over different geographic areas for one taxonomic group will indicate the pattern for all biodiversity. Good evidence for the surrogacy hypothesis is typically claimed when the pattern for the surrogate taxonomic group is congruent with that of some target set of taxa. However, on many occasions such evidence can be explained away as probably arising simply because of a shared bias in the geographic sampling of the surrogate and target taxonomic groups (for review, see Faith 2003). The evidence based on congruence can be explained away as a probable result even without the hypothesis. Based on such evidence, corroboration for the surrogacy hypothesis is low.
Corroboration assessment can provide support for the pattern-models that link objects to units (see section above). Each pattern-model is in effect a surrogate linking objects to underlying units of biodiversity. One approach that has generated controversy is the “ED” method (Faith and Walker, 1996; see above) where the relative biodiversity of a set of localities (objects), at the level of species (units) is inferred by applying a p-median criterion to the environmental gradient pattern. Faith (2011b) reviews evaluations of the ED method where evidence is presented that would not survive a corroboration assessment. He suggests guidelines for achieving a greater role for corroboration assessments in biodiversity surrogates evaluations. These cover three aspects of surrogates testing: experimental design of tests, ongoing corroboration assessment of evidence produced by tests, and accumulation of lessons learned from multiple test studies over time.
The following sections address the potential role of such corroboration assessments in two other areas of biodiversity assessment: phylogenetic inference and species inference (discussion of corroboration assessment in the context of biodiversity monitoring can be found in Downes et al. 2002).
2.1 Phylogenetic hypotheses
The problem of inferring phylogenetic patterns within a taxonomic group from character data raises long-standing philosophical issues. Popperian falsification has been used to argue for the justification of one inference method over others. Cladistic parsimony, which one out of many ways of measuring goodness-of-fit of characters to phylogenetic trees, is characterised as uniquely capturing the idea of falsification (for review, see Faith and Trueman 2001). An alternative perspective is that Popperian corroboration embraces all inference methods in phylogenetics. In this interpretation, the Popperian evidence for a phylogenetic tree hypothesis is a measure of the goodness-of-fit (as defined by any given inference method) of observed character data to that hypothesis. Degree of corroboration of a phylogenetic tree hypothesis is given by improbability of that goodness-of-fit — that is, the difficulty in explaining fit that good by other factors, including elements of chance, that make up our "background knowledge". This reflects the obligation to try to explain-away evidence through identification of some background knowledge that implies that the evidence was probable anyway (Faith and Trueman, 2001; Faith, 2006). The goal of the search is a high probability of the evidence given only background knowledge, even while the desired outcome may be a low probability.
Faith & Trueman (2001; see also Faith 1992) describe corroboration in general terms:
“corroboration assessment requires only the goodness-of-fit or other evidence associated with any phylogenetic method, and background knowledge, which also can take various forms (Faith 1991b, 1992; Faith & Ballard 1994).” Corroboration assessment therefore can support an integrative systematics that uses a wide variety of potential supporting evidence for phylogenetic (or species) hypotheses. Corroboration assessment requires that all supporting evidence be exposed to a skeptical assessment that, in effect, tries to “explain the evidence away‘. Faith et al. (2011) review the arguments that corroboration assessment provides critical examination of evidence, capturing the idea that supposed supporting evidence for an hypothesis is only impressive to the extent that the evidence cannot easily be accounted for by other factors, including chance.
As an illustration of corroboration assessments, Faith et al. (2011) examine recent studies that address the difficulties in determining phylogenetic affinities of Strepsiptera (a highly specialized insect group). Early morphological work had suggested a Strepsiptera-Coleoptera grouping, molecular evidence more recently had dramatically supported a Strepsiptera-Diptera grouping, or Halteria. The good fit of molecular data to the Strepsiptera–Coleoptera tree provides corroboration for that hypothesis because can be judged improbable that evidence so good would result merely from long-branch attraction or poor taxon sampling (Faith et al. 2011). Thus, the supporting evidence is improbable because elements of background knowledge - long-branch attraction, poor taxon sampling, and other factors – cannot easily “explain away” that evidence.
In contrast, supporting evidence for the Strepsiptera-Coleoptera hypothesis is improbable given background knowledge that includes the possible effects of long-branch attraction and poor taxon sampling. Other supporting evidence, however, can be judged probable under background knowledge that includes the possible effects of poor taxon sampling and plausible convergence of thoracic features. These factors can explain-away some character support for the Strepsiptera-Coleoptera clade.
The “Strepsiptera” example illustrates the utility of corroboration assessment. Corroboration assessment can support an inclusive or integrative systematics. However, such assessments are not a well-established part of systematics. Recent work attempts to reconcile this interpretation of corroboration with other, long-standing, philosophical perspectives in systematics on the nature of evidence and hypotheses (for review, see Faith et al. 2011, 2012).
2.2 Species hypotheses
Testing an hypothesis that a set of populations is a single species is important to conservation management. Also, sets of recognised species often form the basis for surrogates for geographic priority setting. Corroboration may play a role in the ongoing debates about the definition of a species and how species status is to be determined. The entry on species discusses the issue of species pluralism — the idea that there is not just one correct species concept. While twenty or more different concepts have been identified (some based on a designated species discovery process), a possible emerging consensus (e.g. see Claridge et al. 1997; Mayden 1997) is that all of these may be unified under an evolutionary lineage concept. This is based on the idea of an evolutionary species, defined as: "a single lineage of ancestor-descendant populations which maintains its identity from other such lineages and which has its own evolutionary tendencies and historical fate" (Wiley 1981, p. 25).
This "primary" concept arguably is compatible with most other proposed species concepts (for discussion, see Mayden 2002). However, a difficulty is that this seems to simply produce many so-called "secondary concepts", corresponding to all the previously proposed ways of detecting and/or defining species. For example, Mayden (1997) refers to these as "operational concepts" that are "tools" for discovering all the different ways to realize the primary concept. The unconstrained use of these tools suggests a "grab-bag" that amounts to reliance as much as ever on expert opinion. An issue therefore is whether a unified species concept can be matched by some unified operational framework for identifying species. Mayden (2002) does claim that there is Popperian "testability" and possible "falsification" for species hypotheses. He argues that the typical process is one in which we do not reject species status if there is no falsification.
Corroboration assessment may be an important missing element in this framework. In much the same role it plays for phylogenetic hypotheses, it can allow many different kinds of evidence (suggested by different secondary concepts), all brought to bear on a single species concept. Evidence for a species hypothesis will be some fit of observations to the hypothesis, and corroboration will depend on the improbability of such goodness-of-fit without the hypothesis (Faith and Trueman 2001; Faith, 2004). This supports a unified species concept as something more than just a shifting of the pluralism problem down one level — the inevitable pluralism now properly reflects the various kinds of evidence that may bear on the same concept. There are no a priori restrictions on the form of the evidence for species hypotheses, but assessment of improbability of evidence is important in avoiding an arbitrary, grab-bag, approach. Further, over time, experience in corroboration assessment in different contexts — for example, for different kinds of organisms — may have lessons about the context-dependent pitfalls of certain kinds of evidence.
Faith (2004; see also Faith & Trueman 2001) consider the nature of corroboration assessment for species hypotheses (p. 12):
“There are no a priori restrictions on the form of the evidence, but much depends on improbability of evidence. A test statement will be some fit of observations to the species hypothesis and corroboration will depend on not being able to explain that without the hypothesis. Any evidence is permitted, but we ask, ‘is it improbable given only the background knowledge?’ For a given species hypothesis, the evidence may have included tests for reproductive isolation, observed morphological or life history differences, molecular differences and so on, and these may vary in the extent to which they survive corroboration assessment… Assessment of the improbability of any supposed evidence is critical. Further, experience over time in corroboration assessment in different contexts, for example for different kinds of organisms, may have lessons about the context-dependent pitfalls of certain kinds of evidence.”
Padial & De La Riva (2010) adopt this corroboration framework in their proposals for “integrative taxonomy” (see also Dayrat 2005). They provide strong arguments for an integrative taxonomy that uses corroboration assessment applied to diverse forms of evidence for species hypotheses. They also discuss clearly the process of explaining-away apparent supporting evidence (p. 752):
“the evidence is not character data, but the goodness-of-fit between the data (of any kind) and the hypothesis, and corroboration is provided when that goodness-of fit is improbable. For example, a hypothesis supported by, say, a coherent pattern of morphological variation with geography, will gain corroboration if we are unable to explain away (i.e. by causes other than lineage divergence) a narrow goodness-of-fit (statistically evaluated) between the observed data and the hypothesis.”
Similar support for a corroboration assessment process, based on attempts to explain-away evidence, is found in the integrative taxonomy framework proposed by Schlick-Steiner et al. (2010) and Steiner et al. (2010). Schlick-Steiner et al. (p. 431) describe a process of seeking explanations for evidence from different “disciplines” (data sources):
“… the evolutionary processes known for each discipline are examined to see if any of these processes could explain why the delimitation hypothesis from a discipline disagrees with that from another discipline…. A few quantitative methods are available for this step, including one to test convergence in morphology and one to distinguish between hybridization
and incomplete lineage sorting in sequence data for large numbers of loci.”
Schlick-Steiner et al. (2010) and Steiner et al. (2010) assess a “failure rate” that amounts to a corroboration assessment of species hypotheses based on improbability of the evidence. Their failure rate evaluates the probability that their identified “evolutionary explanations”, as background knowledge, could account for the supporting evidence for an hypothesis. Schlick-Steiner et al. and Steiner et al. consider the core idea of alternative explanations, and draw on a variety of evidence and background knowledge that captures process knowledge from various disciplines.
The framework explored by Padial & De La Riva and Schlick-Steiner et al. captures many of the key elements of Popperian corroboration, critically including the core idea of explaining-away evidence. These ideas may provide a stronger conceptual framework for integrative taxonomy.
I thank Constanza Pinochet and members of bioGENESIS for many helpful discussions.
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- The Millennium Ecosystem Assessment
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