The review scheme adopted relies on the approach taken by the European Commission within the International Reference Life Cycle Data System defining the “framework and requirements for LCIA models and indicators” (JRC-IES 2011). The following procedure was followed for the methods review: (1) description of relevant cause–effect chains, (2) definition of criteria to evaluate the existing methods, (3) development of sub-criteria specific to freshwater use, and (4) description and review of existing freshwater use assessment methods.
Description of relevant cause–effect chains
Figure 1 depicts the cause–effect chains that link freshwater type and use to potential impacts at the mid- and endpoint level and, ultimately, to the related area of protection of human health, ecosystem quality, and resources (Jolliet et al. 2004). The identified cause–effect chains serve as basis for the development of specific criteria linked to freshwater use. In nature, precipitation water (liquid or solid) is differentiated in three types of water that are interconnected: (1) surface water (river, lake, and sea), (2) groundwater (renewable, shallow, and deep) that is only reached through surface water and soil moisture, and (3) precipitation water stored as soil moisture (also called “green water”) (Falkenmark and Rockstrom 2006). Fossil groundwater compartment is not connected to other freshwater compartments. Freshwater is characterized by less than 1,000 ml/l of dissolved solids (USGS 2012) and encompasses all previously mentioned three types. The impact of freshwater use is related to (1) consumption of one of these water types and (2) withdrawal of one of these water types and release of surface water. Impact of degradative use is considered as withdrawal of surface or groundwater at a given quality followed by release at another quality. However, impact of direct pollutant release in freshwater and resulting cause–effect chains are excluded from the scope of this study, in which there is no value judgment regarding the inclusion of degradative use in considered methods. Related impact assessment approaches are assessed in the ILCD handbook (JRC-IES 2011). Land occupation and transformation as well as rainwater harvesting are a driver for a change in surface water and precipitation water stored as soil moisture. The availability of the latter water type leads to debated potential impacts that are not considered in this work. However, the modification of the hydrological balance following land transformation or occupation is accounted for in the present framework as it corresponds to a modification of the amount of water that reaches the groundwater and surface water (equivalent to a consumption of the corresponding water).
The use of freshwater can generate potential impacts to humans, the ecosystems, and resources. These impacts can be related to water scarcity, water functionality, water ecological value, and water renewability rate and are influenced by the possibility to develop compensation mechanisms. Water scarcity is defined in this work being the water use approaching or exceeding the natural regeneration of water in a given area, e.g., a drainage basin. In this article, water scarcity is considered as a parameter leading to freshwater deprivation by limiting freshwater availability. Freshwater quality is defined as a set of parameters considered to characterize the chemical, physical, and biological properties of freshwater. It is related to a functionality approach, which assesses to which users the freshwater withdrawn and released are functional (Bayart et al. 2010), and can also lead to water deprivation when water of a given quality is not available anymore for specific users. Water ecological value describes the physical relation to and dependency of ecosystems on freshwater (Bayart et al. 2010). Water renewability rate is the natural rate at which the resource is recharged. Compensation mechanisms refer to the use of backup technologies by human users deprived of “functional” freshwater to meet their needs (Boulay et al. 2011b).
The way human health is affected by freshwater use depends on the level of economic development and welfare (Boulay et al. 2011b; Bayart 2008). If this is sufficient, the lack of freshwater will be compensated by the development of backup technologies [such as desalination or the import of water-intensive goods as virtual water (Allan 1996)]. These compensation activities need to be assessed with a new inventory and can, in turn, lead to environmental impacts via other interventions involved in the compensation activities (e.g., climate change impacts caused by energy consumption for desalination). If the level of economic development is not sufficient to cover these costs, freshwater use will lead to water deprivation for primarily three functions which fulfill essential human needs depending on local conditions: domestic use (hygiene and ingestion), agriculture, and aquaculture/fisheries. Industrial functions of freshwater closed to human essential needs (e.g., house building and provision of pharmaceuticals) are not considered in this framework because they are more likely to consider compensation strategies rather than suffering from freshwater deprivation (Boulay et al. 2011b). Water quality degradation leads to water deprivation when it creates a loss of functionality for users who need water at a higher quality level than the released one. Users who are able to use freshwater at that or a lower quality level would not be deprived. The extent of water quality degradation depends on the amount and intensity of chemical, biological, and thermal pollution withdrawn and is related to the sanitation capacity. The withdrawn freshwater represents an adverse impact depriving users from a given amount of water at ambient water quality; the released freshwater (negative LCI flow) results in a burden reduction by making available the same amount of water for users capable to use water at that quality. Current models agree that the way human health is affected by water use depends on the level of economic development and welfare. They acknowledge that under given conditions, water use can lead to deprivation for essential human needs such as agriculture, fisheries, and domestic use and ultimately to malnutrition and spread of diseases. However, there is currently no sufficient information to determine whether freshwater use in a low-income water-stressed region would lower water availability for domestic users or rather only affect other users, e.g., agricultural, fisheries, or industries (Boulay et al. 2011b).
Malnutrition and spread of diseases are interconnected, i.e., malnutrition could, for example, make a person more vulnerable to the spread of diseases, and reciprocally, some enteric diseases could affect the ability to absorb nutrients and thus contribute to malnutrition. Freshwater use ultimately leads to an aggregated impact on human health, generally expressed in disability-adjusted life years (Motoshita et al. 2010a, b; Boulay et al. 2011b; Pfister et al. 2009).
Water use can also affect the ecosystem, for instance, by changes in the river, lake, or wetland flow quantity (e.g., due to surface water withdrawals); changes in the level of the groundwater table (e.g., due to groundwater withdrawal); changes in flow regimes (e.g., due to turbined water use); and loss of freshwater quality. Similarly to human health, degradation corresponds to the consumption of freshwater of a higher quality (with a higher ecological value or ecological functionality) and the release of freshwater of lower quality (with a lower ecological value, thus affecting all the ecological users needing a better water quality, but not the users able to deal with a lower quality).
It should be noted that the latter cause–effect chain is related to the deprivation of freshwater of a given quality and not to the aquatic ecotoxicity, aquatic eutrophication, and aquatic acidification impact of this degradation. The midpoint impacts related to freshwater deprivation, which depend on water scarcity and water quality, eventually lead to species diversity change in aquatic and terrestrial ecosystems. The extent of these changes depends on the ecological value of water in the considered ecosystem. Ultimate impacts on ecosystem quality are commonly expressed in potentially disappeared fraction (PDF) of species on a given surface or volume during a given time (PDF per square meter per year or PDF per cubic meter per year) (van Zelm et al. 2011; Hanafiah et al. 2011).
Milà i Canals et al. (2009) suggest that changes caused by production systems on the amount of rainwater available to other users (ecosystems) through changes in the fractions of rainwater that follow infiltration, evapotranspiration, and runoff should be included as impacts on ecosystem quality. This is closely linked to the impact of land occupation and transformation on green water availability through the variation of stock of water stored as soil moisture available for plant uptake (green water).
Consumption of all freshwater types as well as withdrawal and release of fossil groundwater can respectively lead to overuse of renewable water bodies or exhaustion of nonrenewable fossil groundwater. Overuse of renewable water bodies depends on the water renewability rate. These midpoint impacts affect water flows and funds and ultimately have an effect on the resources stock. This reduction of available water affects other cause–effect chains by increasing local water scarcity. Different approaches exist to characterize the impact on resources encompassing the abiotic depletion potential given in antimony equivalents (Sb-eq) (Milà i Canals et al. 2009) at the midpoint level, and the backup technology concept expressing the resource damage in megajoules (MJ) surplus energy- (Pfister et al. 2009) or exergy-based methods given in megajoules of exergy (MJex) (Boesch et al. 2007) at the endpoint level.
Definition of criteria to evaluate the existing methods
Five scientific (1–5) criteria and one potential stakeholder acceptance (6) criterion based on the ILCD Handbook (JRC-IES 2011) were adopted within this review: (1) completeness of scope; (2) environmental relevance; (3) scientific robustness and certainty; (4) documentation, transparency, and reproducibility; (5) applicability; and (6) degree of potential stakeholder acceptance and suitability for communication in business and policy contexts. They are further described in Table S5 in the electronic supplementary material.
Development of sub-criteria specific to freshwater
In addition to the six criteria mentioned above, sub-criteria specific to freshwater use were added in the criteria “completeness of scope” and “environmental relevance” as described in Table 1. For the former, sub-criteria were needed to identify which areas of protection are considered by the existing methods and which mid- and endpoints are modeled. For the latter, sub-criteria were needed to evaluate the coverage of relevant freshwater-specific cause–effect chains as depicted in Fig. 1. The level of coverage was assessed without weighting the relative importance of different cause–effect chains and related parameters, but rather by exploring how far and with which method this coverage has been performed.
Description and review of existing freshwater use assessment methods
Various methods have been developed to evaluate freshwater use in LCA. Many of them were already published or in the process of being published. All methods addressing freshwater use supported by sufficient documentation to be analyzed, i.e., a draft article, a report, etc., were considered in this paper. Unpublished methods were assessed regarding the latest information available in June 2012. Figure 2 summarizes the reviewed methods and classifies them at the inventory level, water index level, or impact assessment level, distinguishing between mid- and endpoint assessments. It identifies those specifically addressing one area of protection or more comprehensive methods that address more than one area of protection. Databases are called according to the database name and methods according to the name of the developer for academic work, e.g., Boulay (Boulay et al. 2011b) or the industry for methodology developed within a company, e.g., Veolia. A short description of assessed methods is provided in the supporting information.
The inventory section contains both inventory databases and inventory methods. The ecoinvent database (Frischknecht et al. 2004; Ecoinvent 2007) and GaBi database (PE 2011) are widely used databases and contain elementary flows for freshwater withdrawal and turbined water. The WFN database (Water Footprint Network 2011) assesses the inventory consumptive and degradative flows of crops and derived crop products, farm animals, and animal products; biofuels; national consumption and production; as well as trade in crop and animal and industrial products according to the WFN method (Water Footprint Network 2011). Pfister et al.’s database (Pfister et al. 2011) assesses the freshwater consumption for the production of 160 crops. An additional source of data for consumptive and evapotranspirative use can be found for five crops and three livestock products (Hanasaki et al. 2010). The Quantis water database (Quantis 2011) is a database of water uses based on ecoinvent 2.2 developed in the aim of providing industrial stakeholders with datasets required to apply all existing impact assessment methods.
Inventory methods generally suggest concepts for a systematic classification of freshwater elementary flows according to their type (surface water, groundwater, precipitation water stored as soil moisture, whether intake water quality is considered, etc.) without providing respective data. Inventory methods also describe technical water flows such as cooling water and irrigation water. The reviewed inventory methods differ widely in their objective and level of detail. Some focus on defining water categories to allow quality to be considered (Vince 2007; Bayart 2008; Boulay et al. 2011a), and others, on providing inventory tools for organizations (Hoekstra et al. 2011; WBCSD 2010), integrating the effects of direct water use and of land occupation and transformation on water availability in a comprehensive methodology (Milà i Canals et al. 2009), or providing detailed hydrological modeling and classification of freshwater use data in specific sectors (e.g., Australian red meat sector) (Peters et al. 2010). Boulay et al. (2011a) was built on Vince’s (2007) and Bayart’s (2008) methods.
Midpoint assessment methods
Midpoint impact assessment methods give an indicator either common to all areas of protection or specific to a defined area of protection. Methods covering all area of protections giving a single index related to water scarcity include the Swiss ecological scarcity (Frischknecht et al. 2006; Pfister et al. 2009; Ridoutt and Pfister 2010b), Water Impact Index of Veolia, Boulay et al. (2011b) methods, and Water Footprint impact indices (Hoekstra et al. 2011). Area of protection-specific midpoint indicators describe the impact pathway leading to a decrease in freshwater availability for contemporary human users (Bayart 2008), as well as changes in freshwater availability for ecosystems leading to freshwater ecosystem impacts (Milà i Canals et al. 2009) and changes in groundwater availability causing freshwater depletion (Milà i Canals et al. 2009). Milà i Canals et al. (2009) suggest to use different types of water indices (Smakhtin et al. 2004; Falkenmark et al. 1989; Raskin et al. 1997) to assess freshwater ecosystem impacts. Falkenmark et al.’s (1989) index focuses on human use by evaluating the fraction of the total annual runoff available for human use. Raskin et al. (1997) use a water use per resource refined by Smakhtin et al. (2004) by subtracting environmental freshwater requirements from the available resources to derive a water index focused on freshwater resources available for human use.
The overall “blue-green-gray water” footprint concept of Hoekstra et al. (2011) was generally classified as an inventory metric, given that precipitation water stored as soil moisture evapotranspirated by plants (“green water footprint”) and consumptive use of surface and groundwater (“blue water footprint”) represent physical metrics and are not further characterized. However, the gray water footprint can also be evaluated as a midpoint approach as gray water footprint denotes degradative freshwater use by characterizing the chemical pollution in water similar to “the critical dilution volumes approach”,Footnote 1 i.e., an equivalent amount of water needed to dilute an emission below an acceptable threshold. This method thus juxtaposes measurable inventory results of “blue” and green water footprint with a theoretical volume of “gray water” which corresponds to a characterized inventory results. Using the term gray water also creates the problem of having two competing definitions of this term circulating in the water industryFootnote 2 (Henriques and Louis 2011).
Endpoint assessment methods
Endpoint impact assessment methods provide specific indicators for potential damages on the areas of protection of human health (Boulay et al. 2011b; Motoshita et al. 2010b, a; Pfister et al. 2009), ecosystem quality (Hanafiah et al. 2011; Pfister et al. 2009; van Zelm et al. 2011), and resources (Pfister et al. 2009; Boesch et al. 2007).
Other approaches exist to estimate impact on resources that attempt to account for the emergy flows put into place by natural processes to make available a given resource at a given state (Zhang et al. 2010; Rugani et al. 2011) but are not evaluated in this review because they are not specific to the characteristics of freshwater resource. Emergy is defined as the measure of both the work of nature and that of humans in generating products and services, i.e., a record of previously used-up available energy that is a property of the smaller amount of available energy in a transformed product (Odum 1996).
Water indices are originally non-LCA-based indicators that express a measure of human and environmental water needs or of the fraction of resource available to meet these needs. Water indices can be used as characterization factors for midpoint (Raskin et al. 1997; Smakhtin et al. 2004; Falkenmark et al. 1989) and endpoint (Sullivan et al. 2003; Döll 2009) impact assessment methods when applied to freshwater consumptive or degradative use. Such indices can be considered as human use oriented (Gleick 1996; Falkenmark et al. 1989; Ohlsson 2000; Seckler et al. 1998; Sullivan et al. 2003; Döll 2009), ecosystem use oriented (Smakhtin et al. 2004), or cover all three areas of protection (Alcamo et al. 2007; Raskin et al. 1997; Pfister et al. 2009; Frischknecht et al. 2006; Hoekstra et al. 2011; Boulay et al. 2011b). In this work, the terminology “water scarcity index” is related solely to withdrawal-to-availability ratio (Smakhtin et al. 2004; Alcamo et al. 2007; Raskin et al. 1997; Seckler et al. 1998; Pfister et al. 2009; Frischknecht et al. 2006; Bayart et al. submitted) or consumption-to-availability ratio (Boulay et al. 2011b; Hoekstra et al. 2011). Water scarcity indices can be based solely on a measure of water scarcity or include, additionally, a measure of water quality (Boulay et al. 2011b). The details of the implementation of water indices in a LCA context, i.e., the water type to be considered in the inventory phase, needs to be specified in order to make water indices applicable in a method.
Uncertainties are generally large in life cycle impact assessment, especially on the endpoint level, and are yet generally not quantified in most of methods. Only a few authors, i.e., Pfister and Hellweg (2011), reported uncertainties for human health and WSI indicators on watershed and country level.