Assessing the Effects of Climate Change on ...

Human and Ecological Risk Assessment: An International Journal

ISSN: 1080-7039 (Print) 1549-7860 (Online) Journal homepage: http://www.tandfonline.com/loi/bher20

Assessing the Effects of Climate Change on Waterborne Microorganisms: Implications for EU and U.S. Water Policy Rory Coffey , Brian Benham , Leigh-Anne Krometis , Mary Leigh Wolfe & Enda Cummins To cite this article: Rory Coffey , Brian Benham , Leigh-Anne Krometis , Mary Leigh Wolfe & Enda Cummins (2014) Assessing the Effects of Climate Change on Waterborne Microorganisms: Implications for EU and U.S. Water Policy, Human and Ecological Risk Assessment: An International Journal, 20:3, 724-742, DOI: 10.1080/10807039.2013.802583 To link to this article: http://dx.doi.org/10.1080/10807039.2013.802583

Accepted online: 16 May 2013.Published online: 16 May 2014.

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Date: 17 September 2015, At: 06:25

Human and Ecological Risk Assessment, 20: 724–742, 2014 Copyright C Taylor & Francis Group, LLC ISSN: 1080-7039 print / 1549-7860 online DOI: 10.1080/10807039.2013.802583

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HAZARD ASSESSMENT ARTICLES Assessing the Effects of Climate Change on Waterborne Microorganisms: Implications for EU and U.S. Water Policy Rory Coffey,1 Brian Benham,2 Leigh-Anne Krometis,2 Mary Leigh Wolfe,2 and Enda Cummins1 1 UCD School of Biosystems Engineering, Agriculture and Food Science Center, University College Dublin, Belfield, Dublin, Ireland; 2Department of Biological Systems Engineering, Virginia Tech, Blacksburg, VA, USA ABSTRACT Despite advances in water treatment, outbreaks of waterborne diseases still occur in developed regions including the United States and Europe Union (EU). Water quality impairments attributable to elevated concentrations of fecal indicator bacteria, and associated with health risk, are also very common. Research suggests that the impact of such microorganisms on public health may be intensified by the effects of climate change. At present, the major regulatory frameworks in these regions (i.e., the US Clean Water Act [CWA] and the EU Water Framework Directive [WFD]), do not explicitly address risks posed by climate change. This article reviews existing U.S. and EU water quality regulatory legislation for robustness to climate change and suggests watershed modeling approaches to inform additional pollution control measures given the likely impacts on microbial fate and transport. Comprehensive analysis of future climate and water quality scenarios may only be achievable through the use of watershed-scale models. Unless adaptation measures are generated and incorporated into water policy, the potential threat posed to humans from exposure to waterborne pathogens may be amplified. Such adaptation measures will assist in achieving the aims of the EU WFD and US CWA and minimize impacts of climate change on microbial water quality. Key Words:

waterborne microorganisms, climate change, modeling, water policy.

Received 1 November 2012; revised manuscript accepted 13 April 2013. Address correspondence to Rory Coffey, UCD School of Biosystems Engineering, Agriculture and Food Science Center, University College Dublin, Belfield, Dublin 4, Ireland. E-mail: [email protected] 724

Effects of Climate Change on Waterborne Microorganisms


INTRODUCTION Waterborne human infectious diseases associated with feces from humans and animals are becoming a greater concern globally, and they place an enormous potential burden on the human population of many countries (Domingo et al. 2007). Despite advances in preventing waterborne disease, severe outbreaks still occur, for example in the United States (Milwaukee, 1993), Canada (Walkerton, Ontario, 2000), and Europe (e.g., Ireland, 2007) (Coffey et al. 2007; Straub and Chandler 2003). Most outbreaks of waterborne illness in developed regions occur from ingestion of contaminated drinking water or contact with contaminated recreational waters (Smith Jr. and Perdek 2004). In the United States, it has been estimated that almost 1 million illnesses and 1000 deaths occur each year as a result of microbial contamination of drinking water (Warrington 2001). From 1991 to 2002, there were 73 outbreaks of waterborne disease resulting in 415,496 illnesses (Craun et al. 2006). Recent efforts to quantify the burden of disease and health care costs associated with waterborne disease in the United States estimate that over 40,000 hospitalizations and $970 million in treatment costs may be attributable to the five most common etiological agents of waterborne illness (Collier et al. 2012). A study by Kramer et al. (2001) indicated 52 waterborne disease outbreaks leading to 11,777 illnesses based on data from 19 European countries from 1986 to 1996. Between 2000 and 2007 in 14 European countries, there were 354 outbreaks of waterborne diseases related to drinking water, resulting in over 47,617 illnesses (WHO 2009). All of these estimates likely represent a small fraction of all waterborne disease outbreaks that occur as gastroenteritis is notoriously under-reported (Palmer et al. 1997) and illnesses contracted by single individuals (i.e., non-outbreaks) are also typically unreported (Coffey et al. 2007). Surface and groundwater are susceptible to various sources of microbial contamination including agricultural runoff, sewage, wildlife, and domesticated animals (Coffey et al. 2007; Coffey et al. 2010a,b,c; Cho et al. 2011). In agricultural watersheds, the export of microbial contaminants from diffuse sources occurs when precipitation mobilizes microbes on the land surface or within the upper few centimeters of the soil profile, and then transports them to watercourses (Oliver et al. 2007; Coffey et al. 2010a,b,c). Thirty-five percent of impaired rivers and streams in the United States are polluted by fecal coliforms, a non-pathogenic indicator organism associated with health risk, making fecal indicator bacteria the leading cause of water quality impairment for rivers/streams and the second most common pollutant for estuaries nationwide (USEPA 2000). The year 2004 National Water Quality Inventory reported that approximately 246,400 river and stream miles (out of 560,000 assessed miles) in the United States contained unacceptably high levels of fecal indicator bacteria (USEPA 2004a). In Europe, more than half of the 127,000 surface water bodies are reported to be in less than good ecological status or potential, and in need of mitigation and/or restoration measures (EEA 2012). Diffuse microbial pollution sources, in particular from agriculture, are reported to affect most surface water bodies in Europe (EEA 2012). The need for potable water to support growing urban populations raises serious concerns about longer-term sustainable provision of safe water in economically developed countries (Nnane et al. 2011). In turn, current predictions of global climate change include higher average air temperatures, increased precipitation intensity, Hum. Ecol. Risk Assess. Vol. 20, No. 3, 2014

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particularly in the middle and high latitudes (Diffenbaugh et al. 2005), and increased frequency of floods and droughts (Patz et al. 2008), a combination that will result in more and more frequent extremes in seasonal water availability. The impact of both point source (PS) and diffuse, nonpoint source (NPS) pollution may be exacerbated by these possible impacts (Moore et al. 1997). Because of widespread sanitation and effluent regulations on PS, precipitation-driven NPS pollution is the primary source of fecal loading into watercourses in developed countries. In addition, increased in-stream flows due to more frequent and more extreme precipitation events can serve to (re)-mobilize fecal contaminants like bacteria being stored in in-stream sediment deposits (Wilkes et al. 2011). Such scenarios will lead to more peak concentrations of microorganisms in surface water. Drought conditions may also lead to higher peak in-stream concentrations as reduced flow volumes result in less microbial dilution. Reduced in-stream flows can be particularly problematic for PS discharges (e.g., wastewater treatment plants) that may have a constant daily microbial discharge. Peak microbial concentrations strongly determine the infection risk via water consumption (Schijven and de Roda Husman 2005). Since the infective dose of pathogenic bacteria, such as E. Coli O157:H7, is low (Licence et al. 2001), it is very possible that extreme climate events may herald increases in the incidence of waterborne disease (Wilkes et al. 2011; Coffey et al. 2010c; Hofstra 2011; Marcheggiani et al. 2010; Taylor et al. 2011; ten Veldhuis et al. 2010). The risks posed by waterborne pathogens to future communities may be very different than those posed today (Boxall et al. 2009). A shift in health risks may be linked to climate driven land-use change, which may alter both microbial loadings to receiving waters and human exposure pathways (Delpla et al. 2009; Hofstra et al. 2011; Schiedek et al. 2007). Current scientific understanding about the effects of climate change on overall public health risk remains ambiguous, rendering the objectives of regulatory frameworks, such as the US CWA and EU WFD, challenging to achieve. This article reviews existing U.S. and EU ambient water quality regulatory legislation for robustness to climate change and suggests watershed modeling approaches to inform the development of additional pollution control management measures given likely climate change impacts on microbial fate and transport. LEGISLATION Background—EU Water Framework Directive and U.S. Clean Water Act Significant surface water quality degradation during the first half of the 20th century resulted in increased regulations governing surface water management in both Western Europe and the United States (Craun et al. 2006; Kramer et al. 2001). In 2000, the EU WFD 2000/60/EC was introduced with the aim to address the weaknesses of former European water legislation in order to maintain and improve the aquatic environment within the EU (European Commission 2000; Coffey et al. 2007; Gevaert et al. 2009). The WFD seeks to ensure that within the EU by 2015 all water bodies have what is defined as “good status.” The approach used in the WFD represents a radical change from traditional PS effluent-quality regulation toward holistic water quality control where valued water uses include ecosystem maintenance, water supply, recreation, and/or fisheries (Kay et al. 2008). Article 11 of the WFD sets out the type of pollution control measures that must be included in river basin 726

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management plans. These normally include basic measures and, where necessary, supplementary measures. Examples of basic measures linked to the WFD include directives on Bathing Waters (76/160/EEC), Dangerous Substances (76/464/EEC), Groundwater (80/68/EEC), Drinking Water (80/778/EEC), Urban Waste Water Treatment (91/271/EEC), and Nitrates (91/676/EEC). These basic measures complement the WFD’s aims by providing specifics aimed at various pollutants or intended uses. The U.S. Federal Clean Water Act Amendment of 1972 (CWA) requires U.S. states to establish water quality standards for specific designated uses (e.g., bathing, fishing) and to establish surface water body monitoring programs to ensure water quality standard compliance (Keller and Cavallaro 2008). Since implementation of the CWA, the quality of surface water in U.S. water bodies has improved significantly according to U.S. Environmental Protection Agency (USEPA) National Water Quality Inventory reports (Keller and Cavallaro 2008). During the initial implementation of the CWA, monitoring and mitigation focused largely on point sources (i.e., permits and wastewater treatment facilities) (USEPA 2007); however, the CWA’s original mandate of “zero discharge” of pollutants was quickly recognized as unattainable and economically unrealistic (Steele et al. 2008). A series of successful lawsuits against the USEPA and various states in the 1990s forced a new focus on unregulated diffuse (i.e., nonpoint) sources resulting in the development of the total maximum daily load (TMDL) program in 1996 (Radcliffe et al. 2009). When surface water is determined to be impaired (i.e., fails to meet appropriate water quality standards), the CWA requires that a TMDL assessment be undertaken to rectify the impairment. The TMDL provides the basis for establishing ambient water quality-based limits for individual discharges and establishes the allowable pollutant loading for a water body based on the relationship between pollutant sources and water quality conditions (USEPA 1991). Similarities and Differences Between EU and U.S. Water Legislation The common link between the U.S. CWA and efforts supported through the EU WFD is the need for reduced pollutant loads and the means for evaluating mitigation or remediation efforts to address these reductions (Steele et al. 2008). Although significant water quality challenges remain for both jurisdictions, the CWA and the WFD have resulted in improved water quality. The focus has shifted from easily monitored and regulated PS discharges (e.g., end of pipe effluent standards) to pervasive and more recalcitrant NPS pollution in recent decades. Both regions have more rigidly enforced existing, or promulgated new, regulations emphasizing a holistic watershed-scale approach (Daniel et al. 2011; Kramer et al. 2001), including development of environmental objectives considering water quality criteria (e.g., E. coli standards) and use of available programs/tools/resources (e.g., modeling/monitoring) to achieve objectives. A comparison of key components of the WFD and CWA is illustrated in Figure 1. The WFD mirrors many of the same principles in the CWA. For example, U.S. states are responsible for the development of water quality standards, including designated uses, water quality criteria, and anti-degradation policies (Brady 2004), although these are ultimately submitted to the federal-level USEPA for review and approval. Hum. Ecol. Risk Assess. Vol. 20, No. 3, 2014

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Figure 1.

Comparison of the EU Water Framework Directive and U.S. Total Maximum Daily Load program. (Color figure available online.)

Thus, the role of states in the United States is similar to that of EU member countries (Achtleiner et al. 2005). Parallels also exist in terms of objectives, implementation, and approaches (e.g., public consultation or stakeholder engagement, watershedwide management). Despite these similarities in approach, key differences are worthy of note. The WFD places greater emphasis on the desired ecological status as opposed to the pollutant load targets of the TMDL program (Radcliffe et al. 2009). Microbial water quality impairments are the most common motivation for TMDL development studies in the U.S., while nutrients, pesticides, and biochemical oxygen demand (BOD) have received far more attention by the EU’s WFD (Kay et al. 2007). However, it is important to note that TMDLs have been successfully developed for a wide range of pollutants, including nutrients, sediment, pesticides, and dissolved oxygen. To date, there has been relatively scant attention by the EU on the assessment and control of microbial indicator organisms within watersheds (Kay et al. 2008; Coffey et al. 2007). Nonetheless, water policy developments in the EU are progressively concentrating on the need for quantitative information on fecal indicator bacteria (FIB) in watershed systems (Kay et al. 2007). Drinking water supply is identified as the highest beneficial use in the United States for source-water protection (surface and groundwater) and, because of the near certainty of direct human exposure, requires increased focus on the prevention of microbial contamination. In this aspect, the WFD can benefit from lessons learned through the implementation of the CWA in the U.S. to reduce microbial impairment of water sources (Steele et al. 2008; Kay et al. 2008). 728




Within the TMDL framework, coupled watershed-scale hydrology and water quality models are used to relate pollutant loads to in-stream water quality, the contribution from different point and nonpoint sources, and to develop scenarios for achieving the TMDL pollutant load target through percent reductions in identified pollutant sources (Radcliffe et al. 2009). Appropriate model application, in addition to saving significant time and effort on intensive monitoring, allows users to assess alternative pollutant reduction scenarios and forecast TMDL implementation impacts. In the EU, use of water quality models has focused more on the identification of critical source areas, rather than the development of water quality assessment and restoration planning (Radcliffe et al. 2009). Due to the lack of consistent monitoring approaches and the demanding timeframe of the WFD, increasing emphasis is being directed toward using watershed modeling applications to link changes in pollutant sources and water quality with the effects of meteorological drivers (Gevaert et al. 2009; Collins and McGonigle 2008). A prominent and heretofore overlooked issue is that implementation of both the CWA and WFD is taking place during a period of unprecedented global climate change, driven in part by anthropogenic changes to atmospheric composition (Jennings et al. 2009). Neither the EU nor U.S. regulatory structure currently explicitly accounts for the risks posed by imminent climate variation. Control measures to account for changes in climate need to be included in policies to complement existing water quality criteria, contribute to future water quality improvement, and meet the objectives both pieces of legislation. CLIMATE CHANGE AND FUTURE WATERSHED MANAGEMENT Global Climate Model (GCM) outputs indicate it is likely (greater than 90% probability) that high temperatures, extended and more extreme heat waves, and heavy precipitation events will be increasingly frequent as a result of climate change (IPCC 2007; Ficklin et al. 2010). Observational evidence suggests that the hydrologic cycle is becoming more dynamic (IPCC 2007); however, both the EU WFD and U.S. CWA thus far fail to explicitly account for risks posed by climate change despite the fact that future climate variability has obvious implications for long-term implementation and management (Wilby et al. 2006; Ulen and Weyhenmeyer 2007). To date, the primary water-related, climate change focus in the EU and U.S. has been on volume and supply (i.e., water quantity), despite increasing evidence that water quality issues may be of equal significance and concern. As the impact of emerging stressors (e.g., climate change, land use change, new pollutants) on water quality has not received sufficient attention (Hering et al. 2010), the ultimate success of either regulatory strategy in preserving ecological services and protecting public health is fraught with uncertainty. Watershed Modeling Existing operational microbial monitoring is time consuming, expensive, and may not reveal inter-basin or global long-term water quality trends. Widespread use of watershed models, coupled with intermittent monitoring data, can allow planners to extrapolate beyond existing monitoring records. Thus, if used in tandem with GCM precipitation predictions, watershed-scale water quality models offer the potential Hum. Ecol. Risk Assess. Vol. 20, No. 3, 2014

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to simulate the effects of climate change on localized water resources and permit examination of potential future water scenarios. Key to developing the ability to forecast local water quality in the face of global climate change is the need to consider climate change predictions when performing comparatively small watershed-scale water quality modeling. Climate change projections that are typically generated from global-scale models are too coarse for watershed models and often require some level of statistical manipulation (Wood et al. 2004; Zhang et al. 2011; Merritt et al. 2006; Stahl et al. 2008). Statistical downscaling techniques have been widely applied with success in various studies (Stahl et al. 2008; Wang et al. 2006; Widmann et al. 2003; Salathé et al. 2003); details about downscaling techniques are beyond the scope of this article. Several well-tested hydrologic process-based models have been used to assess climate change impacts on hydrology, water availability, land use/plant growth, nutrient loss, erosion, and receiving water concentrations of sediment (Gassman et al. 2007; Daniel et al. 2011; Ficklin et al. 2009; Steele-Dunne et al. 2008; Singh and Gosain 2011). While those same models are capable of predicting receiving water concentrations of microorganisms, thus far, no peer reviewed literature exists on integrated modeling of waterborne microorganisms, hydrology, and climate change effects utilizing downscaled GCM data. In turn, significant challenges remain when modeling microbial fate and transport processes (Jamieson et al. 2004; Kay et al. 2007; Baffaut and Sadeghi 2010; Coffey et al. 2007, 2010a). As predictive modeling of bacteria is notoriously prone to error at present, a refined understanding of the sediment-bacteria re-suspension processes (Jamieson et al. 2004; Coffey et al. 2010a), subsurface transport (Baffaut and Sadeghi 2010), and growth/decay (Kay et al. 2010; Coffey et al. 2010a) is required to improve modeling efforts that integrate the effects of climate change. Advancements incorporating future climate scenarios can provide the basis to develop effective adaptation measures that could be included in policy planning and help to improve future water quality in line with both the U.S. CWA and the EU WFD legislation.

Climate Change Scenarios and Impacts on Microbial Transport Climate change will affect temperatures and precipitation patterns; meteorological factors that will have a direct influence on future land management. Agricultural production practices will have to adjust to these climate variations and intensify to meet the food demands of growing human populations. Combined changes in rainfall, temperature, and associated human land use evolution will have important impacts on fecal bacteria prevalence, dispersion, and exposure routes (Kratt et al. 2010; Boxall et al. 2009; De Roda Husman and Schets 2010; Johnson et al. 2009; Nardone et al. 2010). Changes in land management will also result in higher volumes of greenhouse gas emissions (Nardone et al. 2010; Searchinger et al. 2008; USEPA 2004b) and consequently compound the effects of climate change. Figure 2 illustrates this conceptualization and summarizes the effects of future precipitation, temperature, land use, and agricultural production systems on microbial fate and transport to water sources. A description of the key factors involved is detailed below: 730




Figure 2.

Consequences and impacts of key climate change factors on microbial water quality. (Color figure available online.)

Precipitation Precipitation is a major forcing variable that must be considered when examining the potential impact of climate variation on water quality, particularly with respect to microbial fate and transport. Over half of the waterborne disease outbreaks in the United States in the past 50 years were preceded by heavy rainfall, and outbreaks in general are strongly associated with extreme precipitation and resultant transport of bacteria runoff to receiving water (Curriero et al. 2001). Climate change (i.e., a rise in air temperatures) is likely to speed up the hydrological cycle and increase the frequency of heavy precipitation events, increasing the potential for contaminant transport from both the land surface and deposited sediments (Boxall et al. 2009). A study by Cho et al. (2010) determined that during wet weather, storm wash-off and re-suspension are equally important processes that are responsible for substantial increases of microbial populations. More frequent extreme precipitation events and flooding will result in greater contamination of both surface and groundwater. In addition, flooding could lead to inundation of infrastructure, including drinking water and sewage treatment facilities (Marcheggiani 2010; Taylor et al. 2011; Patz et al. 2008; ten Veldhuis et al. 2010). This has the potential to increase human exposure to a host of waterborne pollutants including pathogenic microorganisms (Kratt et al. 2010). Hum. Ecol. Risk Assess. Vol. 20, No. 3, 2014

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Some climate change scenarios indicate that, in the future, summers will be drier than presently, but that within these generally drier periods a higher proportion of the seasonal rainfall total will fall in fewer more intense precipitation events, increasing the potential risk of summer flooding (Johnson et al. 2009). Winter rainfall totals may also increase, with more rainfall days and greater rainfall intensity due to milder temperatures increasing the water vapor content in the atmosphere. Seasonal changes in precipitation amount and timing are predicted to vary depending on global location. Increases in annual precipitation totals and in the frequency of extreme precipitation events will increase microorganism loading to surface waters in watersheds dominated by NPS inputs (Kratt et al. 2010). In addition, increasing hydrologic extremes may also result in higher microbial concentrations during periods of drought as lower flows result in less dilution of constant PS discharges (e.g., wastewater treatment facilities) (Senhorst and Zwolsman 2005; Johnson et al. 2009). However, although the total microbial load transported from NPS may be higher, the potential does exist for lower in-stream microbial concentrations due to dilution effects from amplified precipitation and flow volume. Temperature A rise in surface water temperatures has been observed since the 1960s in Europe, North America, and Asia (0.2–2◦ C), mainly due to atmospheric warming caused by increases in greenhouse gas emissions from anthropogenic activities (Bates et al. 2008). Surface water temperature is largely dependent on the ambient air temperature and discharge volume. Although local and seasonal differences will occur (e.g., water depth and water flow), the water temperature is expected to continue to rise gradually with the projected increases in air temperature (van Vliet et al. 2011; De Roda Husman and Schets 2010). Higher temperatures may lead to the introduction of new microorganisms, vectors, and/or intermediary hosts (Harrus and Baneth 2005). Enteric bacteria, such as E. coli, are sensitive to warm environmental temperatures (De Roda Husman and Schets 2010); however, inactivation by increased water temperatures may be less relevant for organisms that originate from warm enteric sources, and have more environmentally robust survival structures (Schijven et al. 2003). Thus, higher water temperatures generated by climate change will potentially lead to prolonged survival of such microorganisms in the environment (Hunter 2003). Freeman et al. (2009) suggest that, depending on their growth curve, bacteria may initially grow faster in water with higher temperatures. Schijven and De Roda Husman (2005) point out that recreational bathing in warmer waters (where extended microbial survival is likely to occur) will also play a part in future exposure and is likely to increase the frequency of human enteric infections. Furthermore, increases in water temperature could potentially result in a lengthened transmission season due to sustained survival for microorganisms (Karvonen et al. 2010). Land use Land use evolution, including deforestation and increased urbanization, often contributes to water quality degradation; as climate change influences new patterns of anthropogenic land use, accompanying changes in surface water quality may be an indirect consequence (Delpla et al. 2009). Kratt et al. (2010) suggest that the effects 732




of land use change related to population growth and agricultural intensification may be equal to or greater than those resulting from climate-induced variations in weather patterns. These anthropogenic changes (human and agriculture based) will also result in higher volumes of greenhouse gas emissions that will consequently compound the effects of climate change. Warming is expected to lead to a northward expansion of arable land and more exhaustive production. Of particular significance in Europe will be an increase in forage maize production in the northern latitudes to support livestock production (Olesen et al. 2011). The importance of water management for new grassland species including drainage may, however, be even more important under changed climatic conditions where excessive water logging may occur (Armstrong and Castle 1992). On balance, the projected changes in land use are likely to have adverse impacts on river basin ecology and could lead to an increase in microbial loads within watershed systems. However, it is currently difficult to predict where such increased loads would likely to be more or less prevalent. In addition, predicted population growth will cause increased microbial loading to surface waters through wastewater discharge (Johnson et al. 2009), potentially further exacerbating human health risks from increased microbial loading. Agricultural production systems The majority of waterborne microorganisms are zoonotic agents originating from non-human reservoirs (Sobsey and Pillai 2009). Effects of climate change on animal agriculture may result in increased transport of animal manure to receiving waters, establishing conditions ideal for human exposure and further zoonotic transfer of waterborne microorganisms. Grazing and mixed grain-fed livestock systems, which count on the availability of pastures and farm crops, will be the most affected by climate change. Higher air temperatures will increase thermal stress on unconfined animals and may result in an increased need for indoor housing of livestock during hotter periods to minimize animal stress. Waterlogged pastures from more frequent/intense precipitation events could also result in increased housing of animals. A consequence of increased livestock housing will be higher quantities of stored manure (containing bacteria) that will subsequently be land applied as a soil conditioner and fertilizer; however, impacts from increased applications of manure on microbial transport are uncertain given that bacteria concentrations in stored manure decline over time (Nicholson et al. 2005). Animals exposed to hotter environments also drink 2–3 times more water than those in thermo-neutral conditions (Nardone et al. 2010). This factor may lead to livestock spending increased time in unrestricted surface waters, elevating the probability of direct deposition of fecal bacteria (and possible pathogens) to surface waters. Progressively waterlogged soils could cause compaction and damage due to livestock “trampling” (Falloon and Bretts 2010), causing more annual runoff and increased transport of microorganisms to water sources. The same study also indicates that a longer pasture growing season due to warming will result in greater grassland productivity in Northern and central Europe (Falloon and Bretts 2010). A prolonged growing season will require greater nutrient requirements from fertilizer, potentially in the form of livestock manure and increasing the likelihood of microbial transport Hum. Ecol. Risk Assess. Vol. 20, No. 3, 2014

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to water sources in overland flow. Predicted global population growth will stimulate a need for greater production; recent forecasts suggest that agricultural demand could increase by 25–180% by the 2050s (Weatherhead and Knox 2000; Henriques et al. 2008). More intensive agriculture will yield greater transport of microbes in agricultural runoff and further contamination of water sources. Current literature reflects the ongoing debate about the exact nature of climateinduced changes, for example, some studies report likely reductions in microorganisms in the environment (Poulton et al. 1991; Johnson et al. 2009); others suggest higher microbial loads, and enhanced survival (Hunter 2003; Freeman et al. 2009). Overall, it is anticipated that climate change will lead to increased risk to human health from fecal bacteria (and associated pathogens), particularly for those communities in watersheds dominated by agricultural land use (Boxall et al. 2009), as production systems will likely change dramatically as a result of precipitation extremes and seasonal temperature fluctuations (Nardone et al. 2010; Falloon and Betts; 2010).

DISCUSSION It is evident that future temperature, precipitation patterns, and water availability will impact the levels of microorganisms in both surface and groundwater sources (De Roda Husman and Schets 2010); however, the relative magnitude of this impact is largely unknown, especially when concurrent climate effects on land use and agricultural production systems are considered. Existing water policy in the EU and United States has thus far failed to adequately account for and adapt accordingly to the issues that will arise from global climate change. While current water quality directives and standards are sufficiently stringent to protect water quality, there is a need to assess and develop appropriate adaptation measures to respond to the impacts of climate change and assist in achieving the goals set down by both the EU WFD and U.S. CWA. The WFD does recognize that preparing for climate change will be a major challenge for water management in the EU and is becoming more active in developing climate change adaptation policies. Currently, the EU WFD requires members to review their river basin management plans every six years in order to prepare for and adjust to climate change, but this frequency may not be sufficient given the need for continuing adaptation in response to long-term impacts. Nonetheless, in 2009, the EU issued a white paper and a guidance document on adaptation to climate change in water management to ensure river basin management plans are “climate-proofed” (European Commission 2009). In 2012, the European Commission presented a “Blueprint to Safeguard European Waters,” which includes recommendations on how to ensure that climate change is taken into account (European Commission 2012). In contrast, efforts to integrate the potential effects of climate change within the U.S. Clean Water Act’s TMDL framework have thus far been largely ignored. This may render future evaluation of implemented restoration plans difficult, as Blankenship (2008) reports that anticipated climate changes could negate previous remediation efforts. If water quality is to be improved and maintained in the future, 734




current policies must integrate adaptation measures to prepare for climate variation and adjust to potential scenarios that may arise. Given the current level of scientific understanding, the scale to which climate change will affect water quality and accompanying public and/or ecological health risk is uncertain. Nonetheless, immediate opportunities do exist to quantitatively explore the impact of climate change on fecal bacteria concentrations (and other contaminants) using readily available modeling techniques and climate change projections (Hofstra et al. 2011). The need for more studies coupling climate model outputs and land use changes to investigate impacts on freshwater systems via hydrologic models was emphasized in previous Intergovernmental Panel on Climate Change (IPCC) reports (Kundzewicz et al. 2007). Several authors have also stressed that quantitative tools are required to identify which water bodies are most vulnerable to direct and indirect climate change effects and to guide general watershed management and remediation efforts (Delpha et al. 2009; Jennings et al. 2009; Wilby et al. 2004). Although numerous authors have advocated more extensive use of quantitative water quality models, the need for continuous model refinement cannot be overemphasized. For instance, predictions of microbial concentrations in receiving waters are notoriously plagued by uncertainty. Previous authors have noted that accuracy of simulated hydrology can be particularly compromised during extreme low and high flow events (Kim et al. 2010; Beckers et al. 2009; Benham et al. 2006), patterns that would be expected to increase in frequency during climate change scenarios. Inaccuracies in simulated hydrology for such extreme events would likely compromise prediction of in-stream bacteria and other contaminants. Efforts to accurately model extreme weather conditions (i.e., precipitation and temperature) may be further complicated as a result of uncertainty in simulating the re-suspension of bacteria in bed sediment and bacteria decay/growth in the environment. Thorough assessment and clear representation of uncertainties is essential to advance integrated microbial and climate modeling and would lead to more effective model development (Hoftsra et al. 2011). A systematic approach should be implemented to account for all potential uncertainties/fluctuations (e.g., changes in weather patterns and land use) and assist in deriving adequate scenarios for use in modeling applications (Hoftstra et al. 2011; Boxall et al. 2009). For instance, Johnson and Weaver (2009) detail a possible structure for modeling climate change impacts on water and watershed systems to support management decisions. The framework is based on responding to climate change from the perspective of risk management rather than prediction. It addresses linkages across spatial scales, across temporal scales, and across scientific and management disciplines. A sequenced structure (based on Johnson and Weaver 2009) for modeling climate change and microbial fate/transport in watershed systems is illustrated in Figure 3. It details a stepwise approach to integrated climate change, microbial, and watershed modeling and focuses on simulating relevant futures scenarios, deriving pertinent adaptation measures, and potential risk management strategies. The approach consists of (a) defining the decision context; (b) data collection; (c) assessment of climate data; (d) model development; (e) scenario analysis; (f) risk management and adaptation.


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Figure 3.

Framework for assessing climate change impacts on microbial water quality and watershed systems (adapted from Johnson and Weaver 2009). (Color figure available online.)

In terms of microbial transport, such a framework advocates concentrating modeling efforts on developing appropriate watershed management policies rather than aiming to achieve statistically accurate simulations of microbial fate and transport. Accounting for climate change scenarios in both the U.S. CWA and EU WFD requires improved understanding of meteorological risks together with other predicted stresses including population growth and agricultural patterns (Wilby et al. 2006). Due to current limitations in bacteria fate and transport simulation capabilities, modeling and evaluating climate change from the perspective of risk assessment and watershed management represents the most viable means of developing appropriate strategies to minimize impacts. This concurs with the fourth assessment report of the IPCC, which emphasizes a shift from climate change impact assessment to identifying practical adaptation measures and/or enhancement of adaptive capacity. 736



Simulation of adaptation measures needs to consider realistic, worst case, best case, and likely scenarios. This would provide relevant authorities with the information to make appropriate watershed management decisions, protect water quality from potential impacts, and meet the objectives set down by the EU WFD and U.S. CWA.


CONCLUSIONS It is widely accepted by the scientific community that global climate change will affect levels of waterborne pollutants, including pathogenic microorganisms, in receiving waters, although the exact nature and magnitude of these effects is a matter of continual debate. Concurrent future changes in land use, population increase, and the evolution of agricultural production systems may also contribute to increased microbial contamination of surface waters in the future and, in turn, to increased risk to human health. If water quality is to be improved and maintained in the future, existing EU and U.S. policy must be updated to incorporate control measures aimed at managing potential scenarios and assist in meeting existing water quality criteria. Investigative studies in this area are difficult and comprehensive analysis may only be achievable through the use of watershed-scale water quality modeling applications that possess the capabilities to predict the effects of future climate scenarios efficiently. A systematic approach to model development from the perspective of risk management should be implemented to account for uncertainties and to derive pollution control measures to counteract climate extremes. This approach would provide relevant authorities with the information to make appropriate watershed management decisions and protect water quality. Current water quality standards and directives are sufficient to maintain the quality of water; however, unless realistic adaptation measures are generated and incorporated into water policy, previous remediation efforts to reduce microbiological contamination and improve water quality bodies could prove to be inadequate and the potential threat posed to humans from exposure to waterborne pathogens may be amplified. Such adaptation measures will assist in achieving the aims of the EU WFD and U.S. CWA and minimize impacts of climate change on microbial water quality. ACKNOWLEDGMENTS The authors acknowledge funding under the Marie Curie Fellowship scheme by the Seventh Framework Programme of the European Union.

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