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The Global Freshwater Crisis

Fresh water is essential for life, and thus its provision for agriculture, sanitation, and domestic use is central to meeting many of the Millennium Development Goals and the more-recently proposed sustainable development goals (Griggs et al. 2013; Pahl-Wostl et al. 2013a). However, from a global perspective it is an absolutely limited resource, representing no more than 0.008 % of the volume of water on Earth and covering only about 0.8 % of the global surface area (Mittermeier et al. 2010; see Fig. 17.1).

Fig. 17.1
figure 1

a Approximate quantity and proportionate amounts of all water on earth; b approximate quantity and proportionate amounts of fresh water on earth. Illustration prepared by Stephen Nash

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Fresh water is also a highly threatened resource. A characteristic of the Anthropocene world is a ‘pandemic array’ of human transformations of the global water cycle (Alcamo et al. 2008), including changes in physical, biogeochemical and biological processes. Water scarcity and quality degradation already impact more than 2.5 billion people on Earth, and by 2030 human demand for water is expected to exceed reliable freshwater supply by 40 % (Addams et al. 2009). There is, and will be, every attempt to close this water gap in order to support social and economic growth around the world. Nations have already responded to the threats to human water security by massive investment in water technology and engineered systems (Zehnder et al. 2003; Vörösmarty et al. 2010, 2013). While these engineered solutions might address human water needs, they are not concerned with the biodiversity and ecological function of the systems. Instead they often add to existing threats to biodiversity and ecosystem function. They may involve increased appropriation of surface water flows that are essential for environmental needs, and increased extraction of groundwater resources that are also essential to surface ecosystems and may be non renewable (Taylor et al. 2012; Foster et al. 2013).

Fresh waters are therefore in a state of global crisis; they are perhaps the most imperilled ecosystems on Earth, and inland waters are recognised as hotspots of endangerment (Dudgeon et al. 2006; Darwall et al. 2009; Mittermeier et al. 2010). Nearly every major river has been dammed resulting in the impoundment of over 10,000 km3 of water (Chao 1995; Chao et al. 2008), the equivalent of around five times the volume of the Earth’s rivers, and reservoirs trap more than 25 % of the total sediment load that formerly reached the oceans (Vörösmarty and Sahagian 2000). Around 70 % of available surface water is used annually for agricultural purposes alone (Wallace et al. 2003). Nutrient runoff has created algal blooms and anoxic dead zones. There is a very strong correlation between total phosphorus inputs and phytoplankton production in freshwaters (Anderson et al. 2002; Heisler et al. 2008), and runoff aggravates the formation of coastal dead zones, which have now been reported to affect a total area larger than the United Kingdom (Diaz and Rosenberg, 2008). More than two thirds of our upland watersheds are not protected (Thieme et al. 2010). Wetlands cover about 6 % of the Earth’s surface. Depending on the region, between 30 and 90 % of these wetlands have already been destroyed or are heavily modified (Junk et al. 2013). Climate change will exacerbate the existing threats on wetlands such as land reclamation, pollution, water abstraction, overuse of resources, and facilitate invasion and establishment of exotic species as habitat conditions alter, reflecting (for example) shifts in flow and inundation patterns, increasing temperature and sea level rise.

There are clear signs that freshwater biodiversity is declining rapidly (Dudgeon et al. 2006; Darwall et al. 2009). Population trend data indicate that whereas terrestrial species show declines in the order of 25 % (95 % CL: 13–34 %) since 1970, the equivalent value for freshwater species is 37 % (21–49 %)—nearly one and a half times as high (Loh et al. 2005; and see Fig. 17.2). It should be stressed that these population trend data are based entirely on a selection of water-associated vertebrates, and lack adequate representation from the more species-rich invertebrates (Cardoso et al. 2011; but see also Balian et al. 2008).

Fig. 17.2
figure 2

The Living Planet Index (LPI) tracks the fate of populations of thousands of vertebrate species, just like a stock market index tracks the price of a basket of shares. The global LPI (red line) has declined by 28 % between 1970 and 2008. The global LPI can be split into its components by realm: terrestrial (green line), freshwater (light blue line), and marine (dark blue line). While all components have declined, freshwater has done so much more (37 %) than the marine (22 %) and the terrestrial (25 %) ones

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While existing knowledge is inadequate, at least 10,000–20,000 freshwater species have become extinct within the last century or are currently at risk globally (Strayer 2006; Strayer and Dudgeon 2010). The IUCN Red List of Threatened Species currently only gives partial coverage to the world’s freshwater species, currently listing 23,291, or 18.5 % of all known freshwater species. Accepting that the data may be biased towards inclusion of threatened species present in a region (rather than the more recent trend to provide a comprehensive coverage of all species regardless of the threat; see Darwall et al. 2009, Carrizo et al. 2013), the trends are nevertheless disturbing: 30.1 % of all freshwater species that have been assessed by IUCN are classified as threatened (i.e., ‘Critically Endangered’. ‘Endangered’, or ‘Vulnerable’ according to Red List criteria (IUCN 2013). Amphibians, a primarily freshwater taxon, are the second most threatened group of organisms (after cycads) that have been assessed globally (IUCN 2013; see Text Box 1); but, in intensively-developed regions, over one third of the species in other freshwater taxa are threatened also (e.g. Kottelat and Freyhof 2007; Jelks et al. 2008; Cuttelod et al. 2011; Collen et al. 2014). Although knowledge of freshwater biodiversity is improving (Clausnitzer et al. 2009, 2012; Darwall et al. 2009; Tisseuil et al. 2012; see Text Box 2), information gaps in the tropics (Balian et al. 2008) mean that the overall threat extent may be even greater than currently estimated. The possible extinction of the Yangtze River dolphin, Lipotes vexillifer (Turvey et al. 2007; Smith et al. 2008), which would be the first human-caused extinction of any cetacean, is not only emblematic of the perilous state of freshwater biodiversity, but indicative of our reluctance to effectively address conservation needs. It is a matter of great concern that freshwater biodiversity is largely neglected or insufficiently addressed in almost all water-development projects (Pahl-Wostl, pers. comm.; Vörösmarty et al. 2013); for example, the Bonn declaration that resulted from the Global Water System Project, which gave rise to this volume, mentions biodiversity only implicitly.

The increasing stress on water resources that is associated with increasing population and economic growth of the Anthropocene will likely commit us to further extinctions. To this can be added a substantial (perhaps unquantifiable) extinction debt associated with human actions that have been taken already (Strayer and Dudgeon 2010). The likely consequences of climate change for water availability in rivers do not augur well for biodiversity, at least for some regions (Ngcobo et al. 2013; Reid et al. 2013; Pearce-Kelly et al. 2013; Tedesco et al. 2013). Moreover, and as noted above, likely adaptation measures to be taken by humans to adjust to a warmer world may also be damaging (Palmer et al. 2008), and scenarios for the riverine biota in areas where the human footprint is already pervasive (see Vörösmarty et al. 2010) are especially bleak. Biodiversity loss has been shown to significantly affect the ecological function of ecosystems (Hooper et al. 2012). In the case of freshwater ecosystems this may mean that they have a reduced capacity to provide certain services such as food, nutrient cycling, and water filtration that are essential for supporting human livelihoods and health, beyond the supply of water itself (Horowitz and Finlayson 2011; de Groot et al. 2012; and see below).

Importance of Freshwater Biodiversity

There are at least 126,000 species of freshwater animals and vascular plants; this is estimated as perhaps up to 12 % of all known species on earth, and includes one-third (>18,000 species) of vertebrates, which is far more than would be expected from the limited extent of inland waters (Abramovitz 1996; Dudgeon et al. 2006; Balian et al. 2008, 2010). This total number of species is certainly an underestimate (Balian et al. 2010) since it omits several taxonomic groups that are likely to be rich in freshwater species (e.g., fungi, algae, several ‘protozoan’ taxa). It also does not account for the fact that many new species are being described annually, even in the case of the better known groups such as freshwater fishes and amphibians (for example, since 2005 amphibians are being described at a rate of one new species every 2–3 days; Frost et al. 2006; Reid et al. 2013). Nor does it account for recent losses of species that became extinct before they could be described by scientists. An almost unknown ecosystem type is the vast groundwater body. An estimated 50,000–100,000 stygobiont species, i.e. species that finish their entire life cycle in the subterranean freshwater realm, occur globally (Culver and Holsinger 1992). However, less than 10 % of these species are described up to now (Stoch and Galassi 2010). Ground waters are characterized by a very high proportion of endemic and cryptic species, although there is a major lack of information on their ecology and their functional performance.

Freshwater organisms and their ecosystems are valuable in their own right, but are also vital for providing people with many different goods and services (de Groot et al. 2012; Russi et al. 2013). Russi et al. (2013) have noted that the biodiversity of wetland ecosystems are at the core of the nexus between water, food and energy. However, while biodiversity loss does affect ecosystem function (Hooper et al. 2012; see above), there is limited understanding of this relationship for many ecosystems. It is not known how much biodiversity could be lost without seriously jeopardizing ecosystem functions and services, which makes it very difficult to accurately predict the management needs of freshwater systems under changing environmental pressures (Dudgeon 2010; Stuart and Collen 2013). While much research has yet to be conducted, there is evidence that biodiversity improves water quality (Cardinale 2011) and that the loss of biodiversity impacts human livelihood and well-being (Cardinale et al. 2012). To some extent it may seem obvious that we should expect some relationship between biodiversity and ecosystem functioning as, for example, conservation of fish biodiversity is necessary to maintain a productive fishery (Reid et al. 2013). One possible relationship is that ecosystem function may be enhanced in a near-linear fashion as species richness increases. Alternatively, the loss of species may have no effect on function until some critical threshold, or tipping point, is reached whereupon the remaining species can no longer compensate for loss of the others and complete failure may occur. A third possibility is that functioning may be unaffected by the loss of certain species, but greatly impacted by the loss of others, or even by the order in which they are lost. This last ‘idiosyncratic hypothesis’ holds that the identity of species lost may be more crucial than the number remaining, and there is some evidence that this relationship applies in freshwater ecosystems (e.g. McIntyre et al. 2007; Gessner et al. 2010; Capps and Flecker 2013). Recent findings (e.g., Cardinale 2011; Cardinale et al. 2012), and uncertainty over the form of the relationships between biodiversity and ecosystem functioning (see Dudgeon 2011; Tomimatsu et al. 2013), strongly suggest that it would be prudent to adopt the precautionary principle and minimize further species declines or losses. By the same token, the introduction of non-native species may have marked effects on ecosystem functioning (reviewed by Strayer 2010; see also Capps and Flecker 2013), and should be avoided.

Valuing Freshwater Biodiversity and Ecosystems

Appreciation of the need to protect species and nature for their own sake is taken as axiomatic by many scientists, but is often put aside when it comes to addressing the pressing demands of growing human populations and their need for water security and other necessities (Vörösmarty et al. 2013). One good rationale for halting the degradation and destruction of freshwater systems is that of enlightened self-interest; people rely on rivers lakes and wetlands—not only for water, but the other goods and services that they provide that are of immense value, far beyond the mere economic value of water (Costanza et al. 1997; Russi et al. 2013).

Economic values of inland wetland ecosystem services are typically higher than those of many terrestrial ecosystems. For example, the total economic value of inland wetlands (exclusive of lakes and rivers) was estimated at 25,682 Int.$/ha/year, compared to 5,264 Int.$/ha/year for tropical forests (where ‘Int’ refers to a translation of the original values into US$ values on the basis of Purchasing Power Parity; see de Groot et al. 2012). The non-market services of freshwater ecosystems (e.g., regulating, habitat, and cultural services) represents 94 % of the overall economic value of inland wetlands, and 55 % of the overall economic value of rivers and lakes, according to the data provided by de Groot et al. (2012) (and see Text Box 3 for discussion of a specific example of non-market services). There is now a growing appreciation that sustainable use of all types of wetlands is usually economically more beneficial than conversion to alternative uses if all or most services are taken into account (de Groot et al. 2012). Jenkins et al. (2010) showed that restoration of wetlands in the Mississippi Alluvial Valley can provide a high return on the public investment for the restoration.

This potential economic return from careful management of the natural capital of freshwater ecosystems is important for both regional and global economies. Currently up to 0.75 trillion dollars (750 billion USD) is spent per year to maintain the infrastructure and operating costs of water management around the world, and two-thirds of this expenditure is in America and Europe (Zehnder et al. 2003; Addams et al. 2009; Vörösmarty et al. 2013; Boccaletti, pers. comm). These costs are likely to increase as middle and low income countries start to become more affluent and develop their own infrastructure. Hence, it is important to look beyond the traditional reliance on hard-path infrastructure and to work with nature, and use the natural capital it provides (Palmer 2010; Vörösmarty et al. 2013). The objective of such an approach should be to meet the requirements of regional and global economies while also reducing the intensity of threats to the biodiversity supported by these ecosystems (Totten et al. 2010).

Conservation Gaps (Protected Areas and Their Management)

Despite its ecological, economic, and cultural importance, freshwater biodiversity is evidently not adequately protected by existing conservation actions. Darwall et al. (2011b) compared the distribution of threatened freshwater species (crabs, fishes, molluscs, and odonates) with the distribution of protected areas in Africa. Their results showed that while 84–100 % of the studied species had some part of their range in protected areas, only 50 % or fewer of the species had at least 70 % of their range (mapped to river catchments) contained within a protected area (see red boxes in Table 17.1). Given the high degree of connectivity within freshwater ecosystems, such that impacts can spread rapidly and from areas far outside of the protected part of a species range, this lack of protection leaves freshwater species highly vulnerable.

Table 17.1 Percentage of species within existing protected area networks in Africa
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It has also been shown that freshwater ecosystems are not adequately included in the global network of protected areas (e.g., Allan et al. 2010; Herbert et al. 2010). Globally, almost 70 % of rivers have no protected areas in their upstream catchment (Lehner et al. in prep), and yet upper catchment protection is important because this affects the delivery of water in adequate quantity or quality to downstream habitats. There is, therefore, an important need for careful consideration of optimum placement of protected areas to secure freshwater biodiversity under rapidly environmental alterations.

Holland et al. (2012) describe a methodology for identifying priorities for freshwater protected areas via the development of freshwater Key Biodiversity Areas (KBAs), which has also been used by institutions and funding organisation for planning frameworks (e.g. the Critical Ecosystem Partnership Fund). Freshwater KBAs are defined on presence of threatened and endemic species or ecologically unique assemblages of species (Table 17.2), and are mapped using HydroBASINS (Lehner 2012) which is the best available digital hydrology resource for mapping connectivity within catchments, incorporating river basin boundaries, lakes, and river networks.

Table 17.2 Criteria and thresholds for defining freshwater KBAs, based on Holland et al. (2012)
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The application of these methods to Africa and several parts of Asia (Allen et al. 2010, 2012; Darwall et al. 2011b; Molur et al. 2011) has identified a large number of potential KBAs which may be compared to protected areas to identify gaps in both spatial coverage and management focus. Once these gaps have been identified it is then possible to start developing management plans to address those gaps. However, equally as important as identifying the sites where protected areas should be implemented, is identifying the proper management plans for these locations. Abell et al. (2007) described an integrated approach to selecting and managing freshwater protected areas that first identifies focal sites or habitats that are important for species or communities, then defines critical management zones that would support the integrity of these areas, and subsequently embeds these zones within a wider catchment management scheme that integrates multiple user needs (Fig. 17.3). Such focal sites and crucial management zones would be represented as part of the management approach within a freshwater KBA. The objective is to move beyond protection directed just to the actual sites holding target species, towards protective management of the wider associated catchment.

Fig. 17.3
figure 3

Schematics of proposed freshwater protected area zones as proposed by Abell et al. (2007). a Freshwater focal areas, such as particular river reaches, lakes, headwater streams, or wetlands supporting focal species, populations, or communities. b Critical management zones, like river reaches connecting key habitats or upstream riparian areas, whose integrity will be essential to the function of freshwater focal areas. c A catchment management zone, covering the entire catchment upstream of the most downstream freshwater focal area or critical management zone, and within which integrated catchment management principles would be applied. (Reprinted from Abell et al. (2007). Copyright (2007), with permission from Elsevier)

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Freshwater Management Plans

The importance of well-thought out management structures has been highlighted by several studies (e.g., Broadmedow and Nisbet 2004; Dudgeon et al. 2006; Ramsar Convention Secretariat 2010), and simple, single-factor, ‘rules of thumb’ approaches to management are often unsuccessful. For example, Pittock et al. (2010) outlined the status of five wetlands sites in the Murray Basin, each of which is recognised as an “icon site” for the restoration of ecological health in the basin by the Australian government. Despite such recognition, all of these sites have experienced declines in ecological character. Despite this deterioration, there was limited implementation of any conservation or mitigation measures, and degraded habitat was not compensated nor had it been restored in any way. The most recent government initiatives have been to change flow patterns, but apparently not in a carefully thought-out way, with the result that more stress is placed on some areas in favour of others (Pittock et al. 2010). In addition, a single focus on flows, important as they are, is not a sufficient management response to the array of threats these wetlands face, and a series of multiple-factor initiatives integrated across all five sites would have been more likely to result in conservation gains.

Protected area managers often tend to underestimate the stress on freshwaters in protected areas (Thieme et al. 2012). In addition, even in developed countries, resources are limited: a third of the protected areas in the southeastern United States surveyed by Thieme et al. (2012) lacked any budget for freshwater management or protection, and over half had no staff time allocated to freshwater management activities. At the European level, almost 70 % of rivers fail to achieve “good ecological status” according to the EU Water Framework Directive, and most likely will not meet this status until 2015 or later unless there is significant extra allocation of resources to river protection.

There are a number of specific challenges that face those attempting to manage fresh waters with the aim of conserving biodiversity, while meeting human needs for water. While terrestrial conservation strategies tend to emphasize areas of high habitat quality that can be bounded and protected, this ‘fortress conservation’ approach is not suitable for river segments or lakes embedded in unprotected drainage basins unless the boundaries can be drawn at a catchment scale (see, for example, Dunn 2003). This is hardly ever possible, and the shortcomings inherent in fortress conservation are particularly acute for freshwater biodiversity because protection of a particular component of the biota or habitat, for example in rivers, requires control over the upstream drainage network, the surrounding land and riparian zone, and—in the case of anadromous species and the risk of invasive species—downstream reaches as well. It is a major challenge to reconcile the need for a catchment-scale approach to conservation of freshwater biodiversity when this requires that large areas of land need to be managed in order to protect relatively small water bodies.

Thus all the necessary elements for freshwater management and the conservation of its biodiversity need to be included in water policies. Management of water resources must take account of aquatic biodiversity in and of itself, as well as its contribution to ecosystem functions and the goods and services used by humans, while also establishing monitoring schemes that can underpin adaptive management. Planning conservation initiatives or the activities needed to support them—for example, establishing protected areas and conducting biological inventories (Gaston et al. 2008; BioFresh, 2013)—requires high-quality spatial data on patterns of biodiversity and threat. Unfortunately, prioritization of conservation activities has been largely directed at terrestrial habitats, focusing on primarily terrestrial vertebrates as target species (e.g. Rodrigues et al. 2004). Identification of areas that support particularly high freshwater species richness has lagged behind efforts for the terrestrial realm, and the first attempt at mapping global freshwater ecoregions and hotspots was unveiled relatively recently (Abell et al. 2008). This is an important development because we lack confirmation on whether terrestrial and freshwater hotspots overlap (Strayer and Dudgeon 2010), and the analysis at the scale of river catchments throughout Africa suggests that such overlap is low (Darwall et al. 2011a). In addition, terrestrial vertebrates are poor surrogates for the overall freshwater diversity in a given area (Rodrigues and Brooks 2007).

A recent example of a major conflict among potential users of water is the actual boom in hydropower development, in Europe and globally. Although the utmost principle of the European Water Framework Directive (WFD) is to avoid the deterioration of the status of water bodies, we actually experience an unrestrained development in hydropower production; in particular of small-scale facilities. This rising conflict among different users of water occurs mainly because different directives are responsible for managing the different components of water (e.g., biodiversity conservation, irrigation, navigation, water quality). There is an urgent need to develop synergies among the different users, for the benefit of humans and the ecosystem (Pahl-Wostl et al. 2013b).

Knowledge of the status and condition of the biodiversity present within fresh waters provide an essential basis for making decisions that will allow sustainable management of these ecosystems. Many taxa are good indicators of environmental health. For example, the amphibiotic life cycle of dragonflies (with aquatic larvae and terrestrial adults) and their sensitivity to structural habitat quality, make them well suited for use in evaluating long-term and short-term environmental change in aquatic ecosystems and the associated riparian habitats, which are resources heavily utilized by local communities (Kalkman et al. 2008; see Text Box 4). Amphibians have been used as indicators of the general health of the ecosystem (e.g., Welsh and Ollivier 1998; Rice and Mazzotti 2004). Molluscs—as well as other macro-invertebrates—are sensitive to water quality and flow, and are potentially useful in bio-monitoring programs (Strong et al. 2008); many are also threatened with extinction (Johnson et al. 2013) although global assessments of the conservation status of, for example, freshwater snails are lacking. Global biodiversity databases such as the IUCN Red List of Threatened Species can, through the provision of information on species distributions and their sensitivity to identified threats, help to inform decisions on the potential impact of developments on freshwater ecosystems.

Rockström et al. (2009) defined a set of ‘planetary boundaries’ that describe a safe operating space for humanity. Bogardi et al. (2012, 2013, 2013) noted that in a few decades we may transgress those planetary boundaries for freshwater, indicating that we will have failed as an international community to establish political targets or economic incentives for change. To avoid this, we must develop policies and governance that will protect freshwater ecosystems and ensure the long-term provision of freshwater services to humans (Pahl-Wostl et al. 2013b). An important approach will be to take full account of the “nexus” between water, food and energy, as one of the most fundamental relationships and increasing challenges for society (Bogardi et al. 2012; Lawford et al. 2013a; Russi et al. 2013). While biodiversity, and particularly wetland ecosystems, are at the core of this nexus (Russi et al. 2013), freshwater ecosystems and biodiversity often fail to be considered when this nexus is discussed. Their exclusion may cause a permanent source of conflict because synergies among the various users are not exploited and consensus cannot be achieved. A possible reason for excluding biodiversity and the ecosystem as pari passu partners is the complexity and uncertainty they may add.

Next Steps to Meet Global Conservation and Management Needs

As noted above, substantial gaps in knowledge of global freshwater biodiversity still remain, and considerable research is required to provide baseline data that can be used to inform conservation initiatives and action for this imperilled biota. These data should include satellite and in situ observations, combined with procedures to combine and model these global data sets (Lawford et al. 2013b) (Fig. 17.4).

Fig. 17.4
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Map showing the progress towards completion of Red List assessments for freshwater fishes in different parts of the world

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We need to ensure a better allocation of environmental flows in order to allow for sufficient hydric resources to properly support ecosystem functions while also attending human requirements (Poff and Matthews 2013), and this needs to be tied with research on how climate change will affect those allocations. Modification of flows in some regions is likely to be unavoidable, to meet essential human requirements. When this occurs, the implementation of comprehensive environmental impact assessments with recommendations as to how to mitigate the most deleterious impacts is crucial.

The need for more data is an obvious priority, but conservation biologists must also be ready to make the most of the data that are currently available, and to use these to help landscape managers make appropriate decisions. There are many excellent systems for collating biodiversity data into integrated systems that can support monitoring and measurement of change (Scholes et al. 2012; and see discussion above on the IUCN Red List). Some databases are specifically designed to collate and present ecological information, drawn from multiple data sets, to assist private and public-sector decision-makers in developing ecologically sustainable business and management practices (e.g., see Text Box 5). When developing new analytical tools for evaluating impacts on freshwater biodiversity it will be important to look carefully at the needs of the likely users. In some cases in the past, the relevant users and stakeholders have not been sufficiently engaged during the process of tool development (Morrison et al. 2010).

While awareness of the extent of threats to freshwater biodiversity has grown during the last decade, a great deal more needs to be done in order to conserve it. A major challenge we face is to raise awareness of the tremendous diversity of species living within our freshwater ecosystems, as they remain largely unseen and unvalued. The fact that most freshwater species live in a habitat that very few people explore or appreciate leaves them highly vulnerable to the impacts of the Anthropocene. Many freshwater species, some of which may be truly impressive creatures, such as Pangasianodon gigas, the Giant Mekong Catfish, are heading for extinction yet few people will even notice their passing. As this chapter indicates, it is often the very activities that enhance human well-being and water security which place freshwater species at risk (e.g. Vörösmarty et al. 2010). It remains a huge challenge to manage the Anthropocene global water system in a manner that will meet the water, food and energy needs of people, while allowing for sufficient semblance of natural ecosystem functions to remain in order to sustain biodiversity. For some large, iconic animals it may already be too late to reverse population declines, but it would be a travesty to permit the many freshwater species now recognized as globally threatened to follow path of the Yangtze dolphin into our history books. We already have much of the knowledge and many of the tools we need to protect freshwater biodiversity; we must now demonstrate the will to act.