Treatise on Water Science, Four-Volume Set

EDITOR-IN-CHIEF Peter Wilderer Technische Universitaet Muenchen, Institute for Advanced Study, Munich, Germany

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EDITORS Peter Rogers Harvard School for Engineering and Applied Sciences, Cambridge, MA, USA Stefan Uhlenbrook Department of Water Engineering, UNESCO-IHE, Delft, The Netherlands

Keisuke Hanaki The University of Tokyo, Tokyo, Japan Tom Vereijken European Water Partnership, Grontmij, The Netherlands

Fritz Frimmel Karlsruhe Institute of Technology, Karlsruhe, Germany

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THE IMPORTANCE OF WATER SCIENCE IN A WORLD OF RAPID CHANGE: A PREFACE TO THE TREATISE ON WATER SCIENCE The world in which we live is currently undergoing rapid changes, triggered by outstanding advances in natural sciences, medicine, and technology. As a result, the human population grows to levels never known before. Innovative communication and transportation means permit globalization of economy and urban lifestyle. Cities and city life exert an unprecedented pull. More than half of the world’s population already live in urban settings – the tendency is rising. Cities meet the expectations of immigrants, citizens, and businesses only when served by an appropriate infrastructure. Unfortunately, in many parts of the world cities grow faster than the required infrastructure can be planned, financed, and installed. In many cases, installation of water distribution networks and sewer systems, waterworks, and wastewater treatment plants is often lagging far behind schedule – be it because of the lack of financial resources or because higher priority is given to other infrastructural projects, roads, and highways, for instance. At a larger scale, the water demand of agriculture and industry is growing overproportionally with respect to population size as people shift preference to products requiring particularly high volumes of water during the growth season or during the fabrication process, respectively. Two examples underline this statement – the shift toward meat consumption and the preference of clothing made of cotton fibers. The consumers are often unaware of the water required to raise cattle, swine, and poultry, and to keep cotton fields productive particularly when such fields are located in arid regions as is the case in Uzbekistan, for instance. Although the water demand is increasing, worldwide, the capacity of local water resources is not. It is even decreasing in very many areas of the world, resulting from pollution of water bodies and soil, from over-abstraction of water, and from effects caused by climate change. Water deficits in municipal, industrial, and agricultural settings are the result. In many cases, urban and agricultural areas developed in regions where ab initio freshwater is scarce. Drought situations caused by global warming and climate change amplify the deficit between water demand and water availability. Overabstraction of groundwater to meet the local water demand is a common but unsustainable solution to the problem of water shortage. In areas close to the ocean, over-abstraction causes seawater intrusion and subsequent increase of the salinity of groundwater. Rising sea level caused by melting of shelf ice intensifies the intrusion of seawater not only in aquifers but in estuaries as well. In addition, deterioration of ground- and surface water is caused by excess usage of fertilizers and pesticides, and by uncontrolled dumping of solid and liquid wastes onto land. Aggravation of water deficits in municipal, industrial, and agricultural environments is the result. In the nineteenth and twentieth centuries, health problems and eutrophication caused by pollution of surface- and

groundwater were recognized and solved by legal frameworks and enforcement of regulations, and by investing large amounts of money in the development and implementation of infrastructural concepts and technologies. In high-income countries, design engineers and operators of water distribution and sewer systems, water works, and wastewater treatment plants are well trained, nowadays – a major prerequisite of proper functioning of technical installations. In the mediumand low-income countries, however, responsible management of water resources and effective operation and maintenance of water technology are often foreign words. In the twenty-first century, we are confronted with a comparably much larger and much more complex problem of water management compared to the years past. A new approach to water management and water technology is required in response to the rapid increase at the demand side, and rapid loss of capacity and quality at the supply side. A paradigm shift appears to be urgently necessary. The old paradigm was the answer to the conditions prevailing in the highly industrialized and water-rich regions of the world. Over the past decades, considerable time was available to develop, implement, and upgrade measures capable of solving the specific local and regional problems. This, however, is not the situation we have to deal with today and in the years to come. In future, we have to support people with effective and robust water and wastewater services even if the capacity of the local water resources is critically short. To avoid evolvement of economic and societal instabilities, we are obliged to develop techniques and management concepts which can be implemented in virtually no time. We have to serve people, industry, and agriculture alike while keeping the function of aquatic and terrestrial ecosystems preserved. We need methods which are adjustable to the changing climatic boundary conditions. We need well-educated water professionals in academia, water services, and water authorities who understand the local environmental, economic, and societal framework conditions, to draw appropriate decisions and take responsible action. We need methods which are financially affordable. These methods are to be safe with respect to public health. Moreover, they must guarantee ecosystems to exert their generic life-supporting function. The task to solve the complex issue of water-related problems caused by urbanization and lifestyle changes is challenging because of the speed of change at both the demand and the supply side, and also because of the limitations at the financial side. Business as usual is not a tolerable approach. In the course of a shifting paradigm, we should realize that sectoral approaches (as they were usually taken in the past) are to be overcome. We need to understand that the water quantity and quality issue are inextricably linked to the issue of energy and food supply, and with the issue of land management as well. What we need is a holistic approach. Measures

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are to be taken which permit solution of the energy, water, and food crisis in conjunction with measures which enable restoration of the self-regulating capacity of terrestrial and aquatic ecosystems in harmony with the human demand for land. Scientists and engineers are called to take up the task of problem solving as a challenge and as a chance. Solutions have to be found on the basis of the existing portfolio of knowledge and experience, but open minded with respect to the very local conditions in rapid transition. The Treatise of

Water Science is to be considered as a platform on which innovative research and development may proceed. It summarizes the contemporary state of knowledge in the field of water science and technology and paves the way toward a new horizon. Serving humanity with safe water while keeping the self-regulating capacity of the aquatic ecosystems intact – this has to be our common goal. Peter Wilderer

Preface – Management of Water Resources PP Rogers, Harvard University, Cambridge, MA, USA & 2011 Elsevier B.V. All rights reserved.

1 The Water Crisis 2 Why Studying Water Is So Important 3 Current Global Water Balance 4 Establishing Water Policy 5 Predicting Future Demands for Water 6 Drivers of Socioeconomic Growth 7 Transboundary Conflicts 8 River Basin Politics 9 The Contents of Volume I Acknowledgments References

The Greek philosophers gave us a physical world composed of four elements: land, water, fire, and air. Two thousand five hundred years later we are still focused upon these elements, now conventionally referred to as ecosystem (land), water (water), energy (fire), and atmosphere (air), more aware than ever that these elements are essential for all life on Earth. As human populations have multiplied 700-fold since the ancient Greeks, we are facing major crises with each of these elements. At different stages of human development, each has risen to prominence; control of fire was one of humankind’s earliest and fundamental scientific discoveries. With fire under control, land took on increased salience, humankind’s numbers soared, and we managed to inhabit the entire planet. The other two elements, air and water, were always essential; however, until recently, they were considered so abundant that we would never have to worry about depleting them. However, by the end of the nineteenth century when the globe’s land frontiers were closing and filling in, it was then that we as a species began to notice problems with having contaminated the air and water and that it was becoming difficult to find clean air to breathe safely and unpolluted water to drink. In addition, by the end of the twentieth century we discovered that our profligate use of fossil-fuel energy was in danger of changing the atmosphere in ways that were threatening the survival of our species by causing global warming. We find ourselves now on the threshold of the twenty-first century struggling to survive as a species. As a result, the two most salient global issues now facing humankind are energy and water. How do we manage our survival transition into the twenty-first century and beyond? The intertwined developing resources crises lead us to three critical questions pertaining to the global water situation: 1. Will we have enough water to grow food to feed ourselves in the twenty-first century? By far the largest quantities of freshwater are, and will be, those used in agriculture. Currently, agriculture uses about 4000 km3 of freshwater each year to feed approximately 7 billion people. Even though population growth has slowed down globally, we will still face a population of 9 billion by 2050. The demand for agricultural water is complicated by the fact that as people

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become wealthier their dietary tastes change, moving away from grains toward animal products. The same amount of water that provided food for 10 people subsisting on grains previously now only satisfies the agricultural needs of one person who has moved up the food chain toward animal products. 2. How will we provide water and sanitation for an additional 3 billion urban dwellers? Since 2007 the urban population has exceeded the rural population. This has major implications for sustaining the water and sanitation for cities. For example, China’s urban population is expected to reach 1 billion by 2030. Urbanites typically are wealthier than their rural compatriots, and have radically different water demands, more appliances, washing machines, bathtubs, showers, and flush toilets. Even though the absolute magnitude of their demands is much smaller than the demand of agriculture, water plays an important role in urban public health which cannot be ignored. This is particularly the case in the large cities of Asia and Africa where already there are huge unserved populations demanding water and sanitation services. One study estimates that as much as $22 trillion is needed by 2030 just to meet the demands for water and sanitation services. 3. How should we address the future climate uncertainties? One issue that water engineers always prided themselves on was that they could make robust forecasts of the future, at least good enough to be able to build reservoirs, dams, and embankments that would function well enough under a wide range of actual future outcomes. The very existence of the possibility of climate change seriously challenges our ability to rely upon our forecasts. The shift from using stationary time series as the basis for future forecasting is seriously undermined when faced with the possibility of nonstationarity in the time series. There is a need for creative adaptation strategies that would help avoid rapid collapse of engineered and social systems. This first volume of the Treatise on Water Science has 11 chapters dealing with how to address these questions. It is about managing our water resources and will, hence, focus on water. Without water, life, as we know it, would disappear from the

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planet. Water plays an extremely important role in maintaining a sustainable life on this planet for all species including Homo sapiens. Only those few exotic species that have managed to survive in environments which do not require potable water would survive. We fool ourselves, however, if we focus only on water and ignore the connections of water to the overall use of global resources because each of them has a critical role to play in supporting human life on planet Earth. This is the major concern of modern studies of water resources.

1 The Water Crisis Since 1900 the world’s population has tripled, but its water demand has risen sixfold (FAO, 2009). These two facts have forced the global community to focus on the management of global water resources. The major emphasis has been on making integrated water resources management (IWRM) a reality (Global Water Partnership, 2000). This is the global water crisis whose dimensions are daily beginning to manifest themselves to international agencies, national and local governments, and particularly to individual citizens. Unlike the fear of using up energy, which has occurred very rapidly, sometimes seemingly almost overnight as with the 2007 and early 2008 petroleum price rises, the water crisis is a slower crisis – but a crisis nevertheless. The situation changes imperceptibly from day to day – we do not see doubling of prices over periods of months, but like melting glaciers, it is an inexorable slow burn toward scarcity. The time frame is years rather than months, but every day there are more of us, each making demands on this global resource. Although we can find replacements for fossil fuels to power our cars and heat our homes, there is no alternative to replace water. For most important water uses, such as irrigation and drinking water, there is no substitute. Water, however, is influenced by geophysical and geochemical processes which are highly influenced by climatic change on both the supply and demand sides. On the supply side, drying up lakes and melting glaciers can reduce water availability locally, and on the demand side increased temperatures will increase demands for irrigation of food crops, air conditioning, etc. All of these changes will have to be dealt with under a fairly constant global supply of water. The great irony here is that fossil fuels are usually described as nonrenewable resources – they have a fixed amount and could be exhausted – whereas water is a renewable resource of an essentially fixed amount and is used by everybody on the globe, and cannot be used up in the sense that petroleum can be because it is a renewable resource, but access to it by growing populations overtaxes its availability.

2 Why Studying Water Is So Important Water resources have been studied for millennia. Starting even before the ancient Greek philosophers, Plato (428–348 BC), Aristotle (384–322 BC), and Archimedes (287–212 BC), the Egyptians and the Assyrians had planned, designed, and built major water resource infrastructures throughout the Middle East. The Romans took the Greek concerns about water and

public health and expanded them up to a global scale throughout the Roman Empire. The city of Rome with its 16 major aqueducts was a marvel of both engineering and water management, with a per capita water availability equivalent to current European standards. Over the succeeding centuries, we have theorized, analyzed, and prioritized water in myriad ways. The great scientists and engineers from Renaissance Europe through the end of the nineteenth century, Galileo (AD 1564–1642), da Vinci (AD 1452–1519), Torricelli (AD 1608–47), Pascal (AD 1623–62), Daniel Bernoulli (AD 1700–82), and Darcy (AD 1803–58), just to mention a few of the major contributors, made major breakthroughs which still govern management of water in all its forms today.

3 Current Global Water Balance The International Water Management Institute (IWMI), located in Sri Lanka, has adopted the blue–green water paradigm suggested by Falkenmark and Rockstro¨m (2004), in which the water accounting is done according to whether it is due to evaporation (coded green) or due to the residual surface and groundwater runoff (coded blue). Of the total annual terrestrial rainfall, called the renewable freshwater resources, of 110 000 km3, 56% evaporates by biological processes, forest products, grazing land, and biodiversity; 4.5% is evaporated from rainfed agriculture (crops and livestock); a further 0.6% of the green water is evaporated from irrigated agriculture along with 1.4% from runoff sources (these are called blue water); and an additional 1.3% is evaporated from open water storages from man-made reservoirs and lakes. Cities and industry demand only 0.1% of the total and 36% returns to the ocean. The 110 000 km3 of precipitation on the terrestrial landscape is extremely small in comparison with the total resource base and the amounts evaporated in producing food and fiber. Of course, it should be recalled that the actual withdrawal of water from the ecosystem for cities and industries could be several times larger, but that about 85% of these uses (albeit contaminated) return to the runoff account of the terrestrial system. These issues are explored in greater detail in Volume II of this treatise.

4 Establishing Water Policy It is a commonplace fact that if a resource has little or no value then it will be overused. Therefore, one of the major issues in water resources planning and management is to identify the value of water. The value of water has been pondered by scholars for millennia. Plato observed that ‘‘only what is rare is valuable, and water which is the best of all things y is also the cheapest, as quoted by Hanemann (2006) based on Bowley (1973) from Plato’s Euthydemus. Two thousand years later, in considering the difference between the market price of commodities and their economic value, the eighteenth-century economist Adam Smith compared the value of diamonds and the value of water. In his book The Wealth of Nations (1776), Smith made the distinction between value in use and value in exchange. Water, which has great value in use, often has little value in exchange, whereas diamonds, which have little value

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in use, have enormous value in exchange. Both Plato and Smith pointed out that the market price of an item did not always represent its true value. In order to predict water demand how we value and price water is very important.

5 Predicting Future Demands for Water Predicting future demands for any resource is fraught with difficulties, but the complexity of water and its singular issue of finiteness make it particularly difficult to forecast. For example, how much should we worry about climate change and global warming? Global warming – one of the great scientific debates of the twentieth century – has now the opportunity to become the political debate of the twenty-first century. A few scientists in the nineteenth century warned of the effect on the atmosphere of excessive release of carbon dioxide into the atmosphere due to the burning of fossil (carbon-based) fuels. It was not, however, until the 1950s that serious comprehensive CO2 measurements were made. Since then large amounts of research funds have been expended in the field of climate science. Ultimately, in 1990, the United Nations established the Intergovernmental Panel on Climate Change (IPCC). The IPCC has produced four assessment reports so far dealing with the effects of changing greenhouse gas concentrations in the atmosphere. The IPCC has been incredibly successful in raising the status of the scientific understanding of climate change. The action in the United Nations (UN), however, has now shifted away from scientific research toward political action which will promote mitigation and adaptation strategies that could seriously curtail the increase of CO2 in the atmosphere by the end of the twenty-first century. In its Fourth Assessment Report in 2007, the IPCC identified five key impacts of increasing global average temperature: water, ecosystems, food, coasts, and health. A closer reading of the text shows that many of the most serious impacts on the nonwater areas are, in fact, mediated via water. Therefore, for instance, impacts on food are largely due to hydrological changes; aridity has major impacts on food, ecosystems, and human health. Thinking about the relative issues involved in climate change, Mike Mueller (2007) of the Global Water Partnership (GWP) said, ‘‘if it’s mitigation then the focus is rightfully energy, and if it’s adaptation, it will be water resources!’’ By this, he implied that the bulk of the mitigation strategies deal with handling the use, and development, of new energy resources, and the adaptation strategies will be mostly driven by water concerns. Hence, we need to focus on the water-adaptation strategies, bearing in mind that adaptation for other sectors may include many of the same, or similar, strategies. The pivotal role of water impacts, and hence water’s importance to adaptation, is also stressed in the Stern Review (2006). The local and regional effects on water however are inconsistent among the climate models, often predicting large regional differences in magnitude, variability, and direction of change for the most important hydrology parameter, the precipitation. However, whichever of these models one endorses, there is still a question as to what to do in meeting the future water demands. If we are interested in adaptation to global warming and climate change, it is largely irrelevant

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which of the models we accept, because operationally there will be small differences among the adaptation strategies that one should follow, but major differences would arise if we were following a mitigation strategy. Surprisingly, even accepting the most conservative scenario leaves one in a strikingly similar situation – how to plan for the future under highly uncertain outcomes. The issue boils down to how do we deal with uncertainty in making decisions about water planning and management? Water engineers and hydrologists are supposedly expert at making such forecasts in a very uncertain world. Much of the focus has been on changes in the physical parameters, such as precipitation, stream flows, and evaporation, but rising global populations, coupled with rising incomes, and a concomitant increase in per capita consumption, will inexorably lead to serious consequences for the water resources in many areas of the globe, regardless of what happens to climate change. It is the old Malthusian population/ resources debate from the early 1960s; only now we have India and China moving into the middle classes in a big way. Keyfitz (1976) pointed out many years ago that it is the increasing middle class and their consumption patterns that were going to be the major problem for environmental sustainability. How we can adapt to meet these demands will be the major struggle for the remainder of this century. In planning for the future, we must also be aware of unintended consequences of our actions. One example of this is the current US attempt to mitigate climate change by reducing consumption of carbon-based fossil liquid fuels. During 2007–08, fueled by record crude oil prices, we rushed headlong toward a biomass-based liquid-fuel cycle. Because of their huge demand for cropland, water, and agricultural chemicals, the widespread development of biomass fuels turned out to be a disaster for the poor people of the world whose food budgets could not compete with the middle classes’ love affair with their automobiles. This means that water planners and managers need to worry a great deal about climate change. The consequences of the climate change will become apparent only if the planners work within a holistic framework to ensure that all of the consequences of climate change can manifest themselves.

6 Drivers of Socioeconomic Growth Among the earliest modern commentators on the drivers of socioeconomic growth and decline were Adam Smith, Edward Gibbon, Thomas Malthus, David Ricardo, and Karl Marx. Adam Smith, a Scottish economist, published his Wealth of Nations in 1776, which became the great classic of capitalist economic thinking. Gibbon, an English historian, combed the history of the Roman Empire for clues for these drivers in his Decline and Fall of the Roman Empire (1776–89). Malthus, an English country parson and economist, focused on the relationship between population growth and agricultural productivity in his seminal Essays on Population (1798). Ricardo, an English businessman and economist, focused on the declining economic returns from all forms of production and the increasing costs faced by industry over time. Finally, Karl Marx, a German sociologist and progenitor of Marxism, saw growth

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coming initially from capitalist accumulation and later from the labor of the proletariat. From their writings we see a concern about running out of resources as long ago as the eighteenth and early nineteenth centuries, well before the sixfold increase in population and 40-fold increase in per capita wealth arrived in the early twenty-first century. Malthus and Ricardo were particularly prescient about the roles of population, food, and energy resources. Malthus postulated a geometric rate of growth (like compound interest on a bank deposit) of population and an arithmetic growth (simple interest on a bank account) of land being brought under cultivation and, hence, arithmetic rate of growth of food production. Regardless of where they start, these curves will always intersect after a period of a couple of decades, and Malthus predicted widespread famine or violent conflicts to bring food and population into alignment with each other by ‘misery, war, pestilence, and vice’. Ricardo articulated ‘declining returns’ on investments in resources (coal and iron ore in his time; water, oil, and gas in our time) whereby the best (least-cost) resources are used first, followed by the next best, and so on. Increasing demand for the resource leads to price increases that will continue to rise until the resource becomes too expensive to use. These two nineteenth-century concepts can be used to explain our current water resources crisis and suggest pathways to end the crisis. We see these two concepts at work, for instance, in the case of New Delhi the population growth rate clearly exceeds the rate of possible increase in the water supplies (Malthus). On the other hand, in suburban Los Angeles (LA) as the cheapest sources of water are fully exploited, we see the Ricardo effect of increasing costs at work. When LA was developing in the 1930s, water was available at a reasonable cost, but as more and more people demanded more water the cost of supply was also increasing (the best projects had already been built). Without any technical breakthroughs, this means that the cost per unit of water keeps on increasing as time goes by. Of course, these constraints were also at work over previous centuries, even before they were articulated by Malthus and Ricardo, but Homo sapiens were able to avoid them by expanding our resource base through annexation and colonization, to bring in cheaper resources and food; by finding substitutes for scarce resources; and by improving our technology so that the same amounts of land and resources could be used more efficiently. Examples of these effects are seen in the British response to its nineteenth-century rapid population increases. More food was produced, not in England with its limited land and climate resources, but by Australia and other colonies such as Canada and India. This meant that the agricultural land was no longer a constraint on feeding the increasing population. So, Malthus’ limits and Ricardo’s increasing costs were avoided for the time being; however, since the globe is now pretty much filled up and most of the easiest available water is in use, there are few opportunities to expand the physical supply. The only option available to us now is improving the efficiency of water-use technology, but this is where we run into Ricardo’s increasing cost problem. The real question facing the globe at the start of the twenty-first century is whether we can keep on improving our technologies, or finding cheaper supplies or substitutes. However, just because these adaptations worked well over the past 200 years does not

mean that they will necessarily continue to work. This is the crux of the problem facing global water resources.

7 Transboundary Conflicts In historical times, control of water was the source of major conflicts among users often leading to skirmishes and minor wars. Peter Gleick (2009) tracked the history of water conflicts from 3000 BC to AD 2009. Historically, these have ranged from minor to major conflicts, but in recent times since the 1940s there has been less direct conflict and more attempts to resolve water issues by negotiation. He wrote, ‘‘There has been a lot of discussion about ‘water wars,’ a term that sounds great, but to which I do not subscribe: wars start and are fought for many reasons and while water has often been a target, tool, or objective of violence, it is certainly hard to ascribe the primary reason for any war to water alone’’ (Peter Gleick, 2009). However, the lack of availability and access to water may have been one of the conditions leading to many wars. The lack of access to water can have major impacts on the health and wealth of nations; major occupations, such as fishing and farming, cannot flourish, and the growth of cities will be limited. With the development of nation-states in the sixteenth and seventeenth centuries, the lack of access by downstream users and the control by the upstream populace was firmly established. This meant that, without a treaty, the downstream users were essentially cut off from use of the flowing river. The Industrial Revolution brought serious pollution to the rivers which also impacted the downstream users. The UN’s International Law Commission spent 26 years from 1971 to 1997 drafting the UN Convention on the Law of the Non-Navigational Uses of International Watercourses (1997). As of 2010, it has not yet been ratified by the UN General Assembly by the requisite 35 countries needed for it to come into force. The existence of such a treaty is a good indication of the international community’s intentions to improve the nature of collaboration among the riparians in international and transboundary rivers; however, the inability to ratify the Convention says a great deal about the wishes of upstream countries not to cede sovereignty to a supranational body. Despite the nonexistence of a clear set of laws and treaties, customary international laws have used many principles such as prior consultation, avoidance of significant injury, equitable apportionment, nondiscrimination and nonexclusion, and provision for settlement of disputes embedded in the UN treaty. Moreover, the fact that it has not yet come into force has not hindered the resolution of many smaller water conflicts relying on the common-sense ideas presented above. Moreover, even when ratified, the Convention lacks an effective enforcement mechanism and will thus still rely largely upon the goodwill of upstream parties, or the hegemonic strength of the downstream countries. Water conflicts seem to arise every time a river crosses a boundary. For instance, in the Colorado Basin, despite the existence of the Federal Interstate Colorado Compact, there are still serious water conflicts among the seven US basin states and Mexico. In India, we see similar conflicts regarding the Ganges River, both domestically and internationally, with Nepal and Bangladesh sharing access. Transboundary water

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conflict is one area in which the conflicts among the parties really emphasize the need for clear and transparent rules for cooperation. The present seems to be one of the periods of great interest in international rivers, as experts estimate that there are over 145 countries with at least participation in one or more of the 261 international river basins on the Earth. Over time, there have been as many as 300 river-sharing agreements in Europe since the Treaty of Versailles in 1815. However, almost all of these treaties dealt with regulating in-stream use for navigation, hydropower, fishing, and pollution disposal, all of which did not involve the large-scale diversions of water which now regularly occur with irrigation developments. Large withdrawals typically create very difficult water-allocation problems for the downstream countries and, in history, were typically resolved with violence or threats of violence. In our times, we would rather resort to negotiations than war. The previous period of great concern about transboundary river conflicts was in the 1950s and early 1960s. This period culminated in a successful treaty on the Indus Basin, brokered by the World Bank and signed by India and Pakistan. The accord fueled optimism for resolving other major water conflicts. At that time, basins such as the Ganges–Brahmaputra, the Mekong, and the Nile (and even the tiny Jordan River) were subject to detailed analysis, even to the extent of creating river basin commissions in an attempt to avoid conflict among the parties. Unfortunately, this era of concern came up short. Of these large rivers, only the Indus was eventually successfully developed. Currently, we are experiencing a resurrection of conflict, fueled by the shortage of water caused by rapid development and huge population growth, and possibly global warming. To allocate – or reallocate – the flows of a river is always a political decision. No matter how detailed the technical, economic, and social studies are, hard choices have to be made among the various users who stand to gain and lose from such accords. This is true whether the river is a national river or crosses international borders. However, transboundary rivers imply a level of political decision making that goes beyond local and national interest groups. It requires the ability to negotiate between sovereign nations. All rational planners recognize the value of cooperation on river-sharing issues, from sociocultural terms to trade and economic ones. What is not clear, however, is how to put a value on cooperation; in other words, just how valuable is cooperation?

8 River Basin Politics The problem with purely political decisions is the lack of predictive behavior on which they reside. Thus, many politically inclined decisions have led to a deviation from the scientific–technical-based analysis, which accounts for the quantitative benefits from sharing resources among the coalitions of competing groups. Political considerations are sometimes heavily influenced by noneconomic factors outside the technical analysis. They are often pursued separately and apart from economic objectives, with different personnel and rituals. Political approaches tend to be more descriptive and idiosyncratic than the analogous models in the sciences.

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Ultimately, foreign policy is the most influential determinant of a country’s position on international rivers. Linkage of the river settlement to other outstanding economic and social issues between and among countries is important, as is achieving reciprocity for one’s actions either in the linkage of issues or in sharing benefits that are only achievable through international cooperation. In addition, the climate for agreement is a prime political basis for sharing water resources; this comes about when countries have common or shared technical perception of the problems, networking and contacts at the transgovernmental levels, and the need to be seen as being collaborative as a nation. A wide range of solutions are possible in most negotiations, while the net benefits are not the only consideration; many political issues dominate in shaping the decisions on the locations of the investments which might not necessarily be in the interest of the best technical planning. Managing common property resources is a very difficult endeavor, and the added complexities of transboundary water are no exception. An interesting phenomenon is that river flows have both negative and positive externalities typically working only in one direction, that is, downstream. This pervasive unidirectional feature of water use means that resolution of basin conflicts through mutual control of external effects that work reciprocally is generally ruled out. However, downstream countries can also benefit from some positive external effects of upstream use. Aside from the water allocation problems that arise from the physical sharing of a common resource, there are also many water-quality problems that can arise downstream as an effect of upstream use. Natural processes such as floods and droughts can also cause major downstream effects and are sometimes mistaken for man-made externalities, and thus lead to further mistrust and tensions among the riparian states.

9 The Contents of Volume I In presenting a discussion on water resources, this volume has been constrained by the width of the definitions of what constitutes the field. We have presented just 11 chapters which while they cover a broad range of concerns, they do not, by any means, cover the full range of concerns. We do, however, present materials dealing with the three questions outlined at the start of this chapter: feeding the global population, providing water supply and sanitation to the ever-increasing population, and some approaches to dealing with the huge uncertainties associated with potential global climate change. The first three chapters deal with the broad frameworks of IWRM, governance, and water as an economic good. The chapters attempt to lay the groundwork for dealing with water as a fundamental resource for development. In Chapter 1, Roberto Lenton explores the history and evolution of the concept of IWRM and reports on various assessments and critiques of the concept. In particular, the critics have focused on definitions that tend to be narrowly focused such as a country having a national water policy, or a water law, or the river basin as the focus of planning, or participatory management. Lenton also provides a set of criteria by which IWRM could be viewed in practice. These are a sensible set of criteria

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and if followed would make a major difference in the sustainability of water-sector decisions. In Chapter 2, Edella Schlager takes up the issues of water governance. She emphasizes a bottom-up approach working from the water users up through multiple layers of governance. She emphasizes the fact that the modern approaches to water development and management are mostly based on methodology developed by experts for experts. She claims that the challenge now is to design and fit water governance organizations into complex multiscale and intergovernmental and watershed systems. It is no longer a game for experts controlled and run by experts. In Chapter 3, John Briscoe deals with how to value water in its different uses. His major split is between urban water and water used for irrigation, or the use of water as urban infrastructure and public health and water in its productive mode for food production. In both cases, he describes how to evaluate the direct and indirect benefits associated with water use. He argues that the indirect benefits associated with water use can be as large as, or larger than, the conventionally measured direct benefits. He concludes with an appeal to move away from the conventional formulaic applications of benefit/cost analysis and attempt to identify critical supplementary investments to use more fully the multiplier effects of large infrastructure projects. The next three chapters deal with the practical socioeconomic issues of forecasting the demand for water, the pricing of water and sanitation services, and how to know if interventions in water supply really have the benefits attributed to them. In Chapter 4, Benedykt Dziegielewski and Duanne Baumann point out that credible long-term forecasts of water demand are essential to planning for the long-lived water infrastructure. They show that such forecasts must be based on a high level of disaggregation of demand; the uses of econometric models grounded in economic theories of production and consumption, considerations of potential climate change, and must, above all, provide explicit and plausible assumptions. Dale Whittington in Chapter 5 reviews the role of economic pricing approaches to managing water and sanitation services. In this chapter, he cautions against some of the enthusiasm for investments in social overhead capital expressed by Briscoe in Chapter 4, with the potential for oversubsidization of large projects at the expense of smaller ones. Following up on this theme, Alix Zwane and Michael Kremer, in Chapter 6, examine the evidence whether community-level rural water infrastructure successfully reduces diarrheal disease and conclude that the evidence does not support it. However, from their review of the literature they found evidence that sanitation and hygiene are more important than water quality. The next set of three chapters cover the role of groundwater in providing water resources for many different types of water services, managing water for agriculture, and managing the aquatic ecosystem to provide adequate protection of the environment. In Chapter 7, Lopez-Gunn, Llamas, Garrido, and Sanz assess the development of groundwater over the past half century. They very broadly review the assessment of the total resource available, the economics of groundwater use, institutions and governance of groundwater, and the future sustainability of the resource. They conclude that groundwater may be the most important water source under the more extreme climate-change scenarios, in particular for irrigation in

low-latitude countries. They stress the need for better governance structures for groundwater management and that a much higher level of user participation will be required for sustainable use of the resource. In Chapter 8, Jorge RamirezVallejo makes a comprehensive review of all aspects of managing water for agriculture and concludes that the major challenge in this area is to reverse the serious failure of institutional arrangements at the national and local levels to deal with water correctly. The concluding chapter in this section by Max Findlayson on managing aquatic ecosystems recognizes the interdependence of people and their environment and focuses on the management of water to support the ecosystem and the environment. He concludes with a strong support for the Millennium Ecosystem Assessment as the best approach to managing wetlands and their aquatic systems. He points to the need in the coming decades to address the trade-offs among current and future uses of wetland resources, importantly inbetween agricultural production and aquatic diversity. The final two chapters return to some political and social issues of water resources management. In Chapter 10, David Moreau reviews the problems with implementing ambiguous water policy. He uses the case of the experience in the US of implementing the Clean Water Act especially under the federal system where the states are left to implement national policy. He shows how there are few ambiguities in dealing with point sources of pollution, but many in dealing with nonpoint sources which has led to the Balkanization of the implementation with the individual states essentially ignoring downstream states when setting goals for total maximum daily loads (TMDLs). Fittingly, the volume concludes with a chapter by Casey Brown on risk assessment, risk management in the context of potential climate change. He develops an approach to risk management that attempts to reconcile traditional approaches with our growing knowledge of uncertainty that mark the hydrologic records. He concludes that the water community has focused primarily on the means to reduce the uncertainty related to hydrologic events, but little effort has been devoted to reducing hydrologic risk to society or to communicate risk to promote risk-reducing behavior.

Acknowledgments The editor wishes to thank the following persons who helped in the review of the manuscripts in Volume 1: Chris Basso, John Briscoe, Robert Brumbaugh, Ximing Cai, Torkil JonchClausen, Line Gordon, Chuck Howe, Annette Huber-Lee, Roberto Lenton, Tom Maddock, Suzanne Ogden, Margaret Owens, Cliff Russell, Mike Shapiro, and Richard Vogel.

References Bowley M (1973) Studies in the History of Economic Theory Before 1870. London: Macmillan. Falkenmark M and Rockstro¨m J (2004) Balancing Water for Humans and Nature: A New Approach in Ecohydrology. London: Earthscan. FAO (2009) AQUASTAT 2009: Water and Food Security. http://www.fao.org/nr/water/ aquastat/main/index.stm (accessed September 2010). Gibbon E (1776-89/1989) The History of the Decline and Fall of the Roman Empire, vols. I-VI. New York: St. Martin’s Press.

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Gleick PH (2009) Water brief 4: Water conflict chronology. In: The World’s Water 2008-2009: The Biennial Report on Freshwater Resources, pp. 151. Washington, DC: Island Press. Global Water Partnership (2000) Integrated Water Resources Management, Technical Advisory Committee, Background Paper No. 4. Hanemann WM (2007) The economic concept of water. In: Rogers P, Llamas MR, and Martinez-Contina L (eds.) Water Crisis: Myth or Reality? ch. 4. London: Taylor and Francis. Intergovernmental Panel on Climate Change (2007) IPCC Fourth Assessment Report: Climate Change. Cambridge: Cambridge University Press. http://www.ipcc.ch/ publications_and_data/publications_and_data_reports.htm (accessed September 2010). Keyfitz N (1976) World resources and the world middle class. Scientific American 235(1): 28--35.

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Malthus TR (1798) An Essay on the Principle of Population. London: St. Paul’s Church Yard (printed for J. Johnson). Mueller M (2007) Policy Brief 5: Climate Change Adaptation and Integrated Water Resources Management – An Initial Overview, Global Water Partnership, Technical Advisory Committee. Smith A (1776/1976) An Inquiry into the Nature and Causes of The Wealth of Nations In: Cannan, E (ed.). Chicago: University of Chicago Press. Stern Sir N (2004) The Stern Review on the Economics of Climate Change. Cambridge: Cambridge University Press. United Nations (1997) Convention on the Law of the Non-navigational Uses of International Watercourses, adopted by the General Assembly on 21 May 1997. Not yet in force. http://untreaty.un.org/ilc/summaries/8_3.htm and http:// untreaty.un.org/ilc/texts/instruments/english/conventions/8_3_1997.pdf (accessed September 2010).

1.01 Integrated Water Resources Management R Lenton, The Inspection Panel, The World Bank, Washington, DC, USA & 2011 Elsevier B.V. All rights reserved.

1.01.1 1.01.2 1.01.3 1.01.4 1.01.5 1.01.6 1.01.7

Introduction IWRM at the Watershed Level: Watershed Management IWRM at the Water-Use Systems Level: Agricultural Water Management IWRM at the Water-Use Systems Level: Water Supply and Sanitation Services IWRM at the Basin Level IWRM at the National Level: Policies and Governance IWRM at the Transnational and Global Level: Information Sharing, Cooperation, and Technical and Financial Assistance 1.01.8 IWRM as a Meta-Concept 1.01.9 History and Evolution of the Concept of IWRM 1.01.10 Assessments and Critiques of the Concept of IWRM Acknowledgments References

1.01.1 Introduction This chapter is about the planning and management of water resources. It aims to provide a comprehensive look at the practices and approaches that have come to be known as integrated water-resources management (IWRM). Following GWP Technical Committee (2009), the chapter defines IWRM as the way in which water can be managed to achieve the objectives of sustainable development, and an approach that reflects the need to achieve a balance among economic efficiency, social equity, and environmental sustainability. The chapter begins with several sections that analyze integrated approaches to water-resources management at different levels (from small watersheds to basins, agricultural systems, and national and global policymaking), which have been practiced for some time and about which there is now a considerable body of knowledge. Building on these analyses, IWRM is not to be seen as a single approach but as a wide range of approaches to manage water and related resources – a meta-approach or meta-concept, as it were, which both transcends the various levels of decision making and recognizes the importance of integrating decision making at each level. The chapter then provides some historical perspective on the evolution of the concept of IWRM, concluding with a summary of recent assessments and critiques of the concept. Although the concept of IWRM is applicable to a variety of contexts, this chapter focuses on the management of water in the context of development, that is, on the management of water resources to advance sustainable development and reduce poverty. This means that the chapter examines the management of water through the lens of the major development and environment issues that are currently challenging countries across the world and which are intrinsically interconnected to water resources in one form or another. In particular, it means that the goals of water management addressed in the chapter relate to development, and the kinds of water challenges emphasized in the chapter are those most often found in the context of development – such as how to allocate more water to generate rural livelihoods and grow food in order to reduce income poverty and hunger.

9 9 10 12 13 14 16 17 18 18 19 20

The following six sections provide an overview of integrated approaches to water-resources management at different levels, from small watersheds to basins, agricultural systems, and national and global policymaking. Management at each level shares a common focus on managing water and related resources, to achieve multiple objectives, which has been practiced for some time, and has developed its own discipline, vocabulary, body of knowledge, networks of interested people, and global institutions. Each of the six sections therefore focuses on management at a specific level – explaining the main characteristics of management at this level, summarizing the literature on the subject, describing some of the networks and institutions working at this level, and providing examples of good practices. We begin with the watershed level.

1.01.2 IWRM at the Watershed Level: Watershed Management Perhaps the most salient features of management at the watershed level are (1) the crucial role of land as well as water management and (2) the strong relationships between watershed management and downstream impacts. As a result, in watershed management the need for a close integration of land and watermanagement activities and upstream/downstream considerations is imperative. Reflecting these key features, the World Bank, in Darghouth et al. (2008), has defined watershed management as ‘‘the integrated use of land, vegetation and water in a geographically discrete drainage area for the benefit of its residents, with the objective of protecting or conserving the hydrologic services which the watershed provides and of reducing or avoiding negative downstream or groundwater impacts.’’ In many developing countries, watershed management is a crucial part of rural-development efforts to generate rural livelihoods and increase incomes. The concepts and practice of watershed management have evolved in the last 30 or 40 years, in parallel with the evolution of the concept of IWRM as a whole. Importantly, from an initial emphasis on technology and engineering, driven primarily by downstream

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environmental protection considerations, they have evolved toward more integrated approaches driven by upstream poverty reduction and livelihood generation. As the above discussion suggests, there are several features of watershed management that are important in the IWRM context. First, the management of watersheds involves management of a range of resources, including soil resources, crops, forests, livestock systems, and water in the form of overland flow, streams, and soil moisture. Second, the goals of watershed management are generally economic, social, and environmental in nature, and relate both to communities in the watershed itself (principally increased productivity and incomes) and to downstream (principally flood management and drought mitigation, through more controlled water resources). Third, the instruments involved in watershed management have usually involved a combination of institutional, economic, and environmental measures. Participatory approaches have proved to be particularly valuable. The economic returns from the use of water for productive purposes within watersheds have usually proved to be a hugely important catalyst and incentive for collective environmental-preservation activities. A crucial feature of many watershed management efforts, as illustrated by the Sukhomajri experience (see Box 1), is the interplay among environmental, economic, and institutional approaches. Darghouth et al. (2008) have noted that ‘‘where communities could see the economic benefits and were empowered, they were willing to invest in long term conservation.’’ By now, there is a significant body of knowledge and experience in watershed management, which draws on both the range of watershed management experiences in many countries and a growing body of research and evaluation carried

out by institutions such as the World Bank, the International Water Management Institute (IWMI), the World Agroforestry Centre, and the Food and Agriculture Organization (FAO). These and other institutions have established some important watershed management programs and networks. The World Agroforestry Centre, for example, has established the Rewarding Upland Poor for Environmental Services (RUPES) program in Southeast Asia, which aims to develop ‘‘mechanisms for rewarding the upland poor in Asia for the environmental services they provide.’’ FAO has also been active in the subject area, in terms of both publications and networks, playing a key role, for example, in the Latin American Technical Cooperation Network on Watershed Management (RDLACH), which was created in 1980 to promote watershed management in Latin America and the Caribbean. The World Bank has been an active player in watershed management, providing finance for important initiatives and evaluating and drawing lessons from many watershed-management programs in different parts of the world (see e.g., World Bank, 2003, 2004a, 2004b, 2004c, 2005). A recent World Bank discussion paper (Darghouth et al., 2008) summarizes these experiences and lessons learned. The IWMI has carried out important research in this area, as illustrated by Sharma et al. (2005). The Comprehensive Assessment of Water Management in Agriculture (CA), which was spearheaded by IWMI, has also sponsored research on watershed management. One study under the CA (Joshi et al., 2005) carried out an in-depth analysis of the impacts of watershed management programs in India, evaluating 311 case studies of watershed programs in terms of economic efficiency, equity, and sustainability, and concluding that these programs yielded an average internal rate of return of 22%. Other

Box 1 Watershed Management in Sukhomajri, India. From Lenton R and Walkuski C (2009) A watershed in watershed management: The Sukhomajri experience. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development. London and Sterling, VA: Earthscan. Sukhomajri is a small village of about 450 people on the edge of the Shivalik mountain range near Chandigarh in India. By the 1970s, over a century of heavy logging in the area and the overgrazing of cattle, sheep, and goats in open forest lands had severely degraded the area surrounding Sukhomajri. Its people were impoverished and survived primarily by raising rainfed crops and keeping goats that foraged in the denuded hills. The Sukhomajri program came about because the citizens of Chandigarh, whose Sukhna Lake had lost nearly 70% of its storage capacity due to siltation by the early 1970s, asked the nearby Central Soil and Water Conservation Research and Training Institute (CSWCRTI) for assistance in solving this problem. The Institute soon found that the lake’s siltation was caused by soil erosion in the hills in and around Sukhomajri, and developed a program to improve soil and water conservation in the watershed in consultation with the local community. While initially the people of Sukhomajri were not very interested in a project principally designed to protect Chandigarh’s lake, their attitude changed dramatically when the Institute built a small dam to control runoff and the villagers found they had a reliable source of water relatively close at hand. A cooperative effort began, focused on community participation in decision making and management, incentives for villagers to graze their animals outside of the watershed, and – crucially – an equitable system of water allocation that would benefit all villagers equally, with water rights granted to all villagers whether or not they owned land. Over time, the villagers organized and formed a water-users association and later the Hill Resource Management Society, to manage and distribute irrigation water. All these actions underscore the Sukhomajri program’s balance of economic efficiency, social equity, and environmental considerations. Over time, the impact of the Sukhomajri program has been considerable. Annual household incomes, for example, rose from around US$230 in 1979 to about US$1360 in the 2000s – more than double the per capita income of the state of Haryana, which itself is one of the highest in India. In addition, tree density in the area rose a 100-fold from 13 ha 1 to around 1300 ha 1 between 1979 and 1995, underscoring its environmental regeneration benefits. Socially, the Sukhomajri program has had important equity impacts because of the focus on landless people and equal distribution of irrigation water. More broadly, Sukhomajri’s approach to community-integrated watershed management has become a model for watershed-development programs elsewhere in India. Importantly, watershed management in Sukhomajri has undergone many changes, not all of which have been positive, since the program began. For example, the check dam led to a rise in groundwater levels, which led an increasing number of villagers to build shallow tubewells and use these tubewells rather than the check dam for irrigation water; this development lowered the incentives for limiting watershed grazing and participating in communal activities. In addition, changes in taxation led to significant declines in the income for the Hill Resource Management Society. This fall in revenues, coupled with the lower incentives for villagers to participate in watershed conservation, led to silting dams and deteriorating pipelines.

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important works on watershed management include Farrington et al. (1999), Heathcote (1998), and Sharma et al. (2005). An excellent example of watershed management in the context of IWRM is the Sukhomajri program in Northwest India, which has been extensively documented and analyzed by Seckler (1986), Agarwal and Narain (1999), Kerr (2002), Khurana (2005), CSE (1994), CSE (1998), CSE (2002), and CSE (2007). This case is summarized in Box 1.

1.01.3 IWRM at the Water-Use Systems Level: Agricultural Water Management The key feature of IWRM at the agricultural-systems level, commonly referred to as agricultural water management (AWM), is that it is defined by a particular type of water use – agriculture – which in most countries is by far the largest consumer of water. AWM thus has a common purpose, generally defined in terms of increasing and sustaining agricultural production. Management itself takes place at several levels, from a single farmer’s field to small farmer-managed systems to large publicly operated irrigation systems, usually with policy and other support at higher levels as well. Reflecting these features, AWM may be viewed as encompassing the range of structural and nonstructural measures at various levels to harness, control, and manage surface water, groundwater, and rainwater to improve and sustain agricultural production. (This definition draws on the definition used in World Bank (2006), but has been broadened to encompass both large-scale and small-scale, structural and nonstructural interventions at a variety of levels.) Structural measures include combinations of irrigation, drainage, and flood control, water conservation and storage, on-farm water management, and soil-moisture conservation, while nonstructural measures include institutions and policies to improve physical and financial sustainability, and user operation and management . AWM involves increasing access to reliable and affordable water supplies, improving management of rainwater, soil moisture, and supplemental irrigation, finding ways to gain higher yields and value from the same water amounts, and enhancing management of the resource as a whole. The interventions involved vary significantly from level to level. While at the farm level, AWM interventions might involve investments in irrigation or soil management, at the farmermanaged irrigation system level, they also include community mobilization, and at the large-system level, the operation of canals and the governance of resources. Decision making at national policy levels is also a fundamentally important aspect of AWM, as discussed later in this section. AWM, like water-resources management as a whole, is best viewed as an integrated, holistic process. Indeed, AWM is an aspect of water-resources management, and should be understood as such. Water has several characteristics that impact on its management and use in agriculture. For example, water has many competing uses outside of agriculture, and it is required in relatively large quantities to produce yield increases, but is heavy, bulky, and costly to transport in comparison to other inputs.

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While earlier it was noted that AWM is defined by the common objective of increasing and sustaining agricultural production, in fact, AWM generally has broader objectives beyond enhancing agricultural productivity, such as improvements in livelihoods and incomes, reductions in risk, and long-term sustainability of the resource. Increasing the numbers of small holders with access to reliable and affordable water provides more regular employment and livelihood opportunities to landless people as well as small holders, and enhances the prospects of ensuring access to domestic water supply and sanitation in rural areas. Efficient AWM creates opportunities for farmers to improve livelihoods, leveraging investment in other productive inputs such as improved seed and fertilizer, while also helping to ensure longterm sustainability of both surface water and groundwater resources. Finally, good AWM aims to reduce the risks that farmers and countries experience from variable rainfall, which can have a powerful impact on growth. All this reinforces the notion that efficient AWM, like IWRM as a whole, generally has a broad range of economic, social, and environmental goals. Importantly, AWM generally involves the integrated management of both blue water – water withdrawn from rivers, reservoirs, lakes, or aquifers for irrigation purposes – and green water – rainfall stored in soil moisture. While blue water is visible and its role in irrigated agriculture is clearly understood, green water and its crucial role in rainfed agriculture often goes unrecognized. Green water management measures to improve agricultural productivity can encompass soil management, crop choices and practices, and water capture. Blue water irrigation management, on the other hand, can involve water storage, lift, transportation, delivery, application, and reuse at various levels. However, average yields from rainfed agriculture using green water are much lower than those from irrigated agriculture using blue water, and, as a result, only half of the world’s food is produced under rainfed conditions practiced by the majority of the world’s farmers. Better green water management can reach relatively large numbers of farmers at relatively low cost, but the productivity gains are relatively small; improved blue water irrigation management, on the other hand, can achieve higher productivity gains, but reach relatively smaller numbers of farmers and with a relatively high cost per farmer. Within blue water, AWM generally involves the management of both groundwater and surface water. Like green water, groundwater is less visible and frequently overlooked, but enables farmers to exercise much greater control over the amount and timing of water applied to their crops than those forms of irrigation that depend on unreliable surface-water supplies, and is an effective way for small holders to improve crop production. In South Asia and North China, in particular, groundwater irrigation over the last 30 or 40 years has grown considerably, and played an ever-greater role in efforts in these regions to improve productivity, food security, livelihoods, and incomes. Indeed, according to the CA (CA, 2007), the rapid growth of groundwater irrigation in South Asia and the North China plains between 1970 and 1995 was at the heart of the agrarian boom in these two regions. Nevertheless, this boom has come at the cost of decreasing water tables that threaten its long-term sustainability (Shah, 2009).

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There are numerous examples of best practices in AWM that have adopted an integrated approach. One such example is summarized in Box 2. While the physical management of water for agriculture takes place at the agricultural-systems level, decision making at national policy levels is a fundamentally important aspect of AWM. Some of these policy issues relate specifically to irrigation, such as the balance of large-scale and small-scale irrigation, the balance of government-led and community-based interventions, and the role of the private sector. While the institutional, technological, and environmental problems that need to be tackled to improve irrigation performance are usually system specific, the use of economic instruments to increase the efficiency of irrigation is usually a matter of national policy. Other national policy issues relate to agricultural water as a whole. In many countries, for example, a major policy question is the scope for growth in food production in rainfed and irrigated agriculture, from which combination of the two should future food production lie, and what are the wateravailability implications of each of these strategies. A related topic is the extent to which trade, as against domestic food production, should be used as a strategy for sustainable food security, taking into account not only the virtual water transfers that are embodied in food trade but also the risks and uncertainties associated with trade in food supplies and the role of domestic food production in providing a source of

income and livelihoods to rural populations. A third topic is the water and energy nexus. In countries in which both energy and water resources are scarce, a key policy question is whether scarce water supplies should be allocated for biofuels or for hydropower, and if so what should be the crop choice. There is a vast literature on AWM. While much of it is narrowly focused on specific technical issues of irrigation, there is a growing literature that takes an integrated approach and addresses some of the broader policy issues identified above. A recent review commissioned by the Technical Committee of the Global Water Partnership (GWP) (GWP Technical Committee (2009)) showed the wide range of available literature on subjects such as water–food interactions, including trade in virtual water, biofuels and their implications for water, and water efficiency and productivity, in the works of Rosegrant et al. (2002); de Fraiture et al. (2008), and Hellegers et al. (2008). The most recent comprehensive review of AWM, however, is the CA, a recent multi-institutional assessment of the current state of knowledge on how to manage water resources for agriculture. This assessment is summarized in Box 3. Several global institutions have played a major role in supporting national efforts to improve AWM. The World Bank has long been a major source of financial and technical assistance in AWM, investing some US$13.2 billion in 56 countries in the 10-year period between 1994 and 2004 alone (World Bank, 2006). FAO has had an active program in AWM

Box 2 Irrigation reform in Mali. From Barry B, Namara R, and Bahri A (2009) Better rural livelihoods through improved irrigation management: Office du Niger (Mali). In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development. London and Sterling, VA: Earthscan. The Office du Niger in Mali was formed in the 1930s as a centralized public enterprise to produce irrigated cotton and rice. Starting in the 1990, the Office du Niger was significantly revamped through a process involving measures such as physical rehabilitation and modernization, farming-systems intensification, improvements in land-tenure security, the creation of pro-farmer support services, the establishment of farmer organizations, and the introduction of innovations in agricultural technology, along with macroeconomic and cereal market reforms. All these have led to dramatic gains in rice production and farm incomes as well as reductions in rural poverty. The case shows that changing agricultural water management requires a supportive macro-policy environment, appropriate institutional changes, and infrastructural investments. Equally important, it shows that those reforms may need to precede improvements in water management. Moreover, in aid-dependent low-income countries, reform cannot occur unless both government and donors concur on the need for change. Finally, the case drives home that improving water management is a continuing process; gains to date in economic efficiency and (to a lesser extent) equity in the Office du Niger now need to be matched by improvements in environmental sustainability.

Box 3

The Comprehensive Assessment of Water Management in Agriculture. From http://www.iwmi.cgiar.org.

The Comprehensive Assessment of Water Management in Agriculture (CA, 2007) was a critical evaluation of the benefits, costs, and impacts of the past 50 years of water development, the water management challenges communities face today, and the solutions people have developed around the world, aimed at assessing the current state of knowledge and stimulating ideas on how to manage water resources to meet the growing needs for agricultural products, to help reduce poverty and food insecurity, and to contribute to environmental sustainability. The assessment was produced by a broad partnership of practitioners, researchers, and policymakers and organized through the CGIAR’s Systemwide Initiative on Water Management (SWIM). SWIM was convened by the International Water Management Institute (IWMI), which initiated the process and provided a secretariat to facilitate the work. The CA’s scope was water management in agriculture, including fisheries and livestock, and the full spectrum of crop production from soil tillage through supplemental irrigation and water harvesting to full irrigation in a sustainable environment context. The review covered a range of topics, from managing water in rainfed agriculture and groundwater use in agriculture to agricultural use of marginal-quality water resources and integrating water and livestock development. The assessment was originally framed by 10 questions, later expanded as interest grew, and included the overarching question: How can water in agriculture be developed and managed to help end poverty and hunger, ensure environmentally sustainable practices, and find the right balance between food and environmental security? Some of the questions addressed by the CA included: What are the options and their consequences for improving water productivity in agriculture? What are the options for better management of rainwater to support rural livelihoods, food production, and land rehabilitation in water-scarce areas? What are the options for integrated water resources management in basins and catchments? What policy and institutional frameworks are appropriate under various conditions for managing water to meet the goals of food and environmental security?

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for several decades, focusing on knowledge generation and dissemination, policy and technical advice, and preparation, investment, and implementation of projects. On the research and development side, the IWMI has played a major role in shaping thinking and action on AWM. IWMI’s evolution from a narrow blue-water focus on the management of irrigation at the systems level to its current broad mission to ‘‘improve the management of land and water resources for food, livelihoods and the environment,’’ which encompasses work at all levels, mirrors the evolution of the discipline of AWM as a whole.

1.01.4 IWRM at the Water-Use Systems Level: Water Supply and Sanitation Services While agriculture consumes the lion’s share of the world’s water resources, AWM is not the only example of a level of water management defined by a particular type of water use. Water supply and sanitation (WSS) can also be considered as a type of water management defined by its use, with a common purpose generally understood in development circles as increasing sustainable access to basic sanitation and safe water supplies for domestic purposes. As with AWM, management takes place at several levels, from the household to large cities, with policy and other support at higher levels as well. Unlike AWM, however, the amounts of water used for WSS are relatively small. The key feature of WSS is thus that the management of resources other than water – financial, human, and institutional resources in particular – is often more important than the physical management of water itself. Despite or perhaps because of this key difference, water and sanitation-services management has followed a course similar to AWM, in that it has evolved from a somewhat narrow focus on technologies to a much broader understanding of the political, institutional, and financial dimensions of improving access to water and sanitation at all levels. As a result, it is generally agreed today that increasing access to water and sanitation services by the unserved will require a broad integrated approach, at many levels, involving increased political commitment, institutional and technological innovation, and increased financial allocations, with attention not only to the supply side but also to the demand side (Lenton et al., 2005). Several global institutions have played a major role in shaping the thinking in this field as well as supporting national efforts to improve access to water and sanitation services. These include the World Bank (both through its investment program and its WSS program), the World Health Organization, and the United Nations International Children’s Education Fund (UNICEF), the Water Supply and Sanitation Collaborative Council, and several major international nongovernmental organizations, such as WaterAid.

1.01.5 IWRM at the Basin Level The key feature of water management at the basin level is that the basin is the basic unit for integrating the supply side of water-resources management, that is, for integration within the natural system. This means that the basin is well suited for

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integrated management of land and water management, green and blue water, surface and groundwater management, water quantity and water quality, spatial and temporal variability, and upstream and downstream interests. The chapter on freshwater in Agenda 21, the key document resulting from the 1992 Earth Summit (UNCED, 1992), states that ‘‘Integrated water resources management, including the integration of land- and water-related aspects, should be carried out at the level of the catchment basin or sub-basin.’’ Likewise, Article 26 of the Johannesburg Plan for Implementation, the key outcome of the 2002 World Summit for Sustainable Development in Johannesburg (WSSD, 2002), states that ‘‘the river (or water) basin should be used as the basic unit for integrating management.’’ Water management at the basin level has thus become the central focus of much of the advances in thinking about IWRM. Nevertheless, while basin boundaries provide a useful way of delimiting the supply side of the equation, they are not necessarily the best means to integrate the demand side, especially since basin boundaries usually do not coincide with political or administrative boundaries. Integrating natural and human systems therefore generally requires work at other levels beyond the basin. Basins can be characterized by the degree of pressure placed on a basin’s water resources, that is, by the relationship between the requirement for freshwater resources and its availability in the basin, taking into account both quantity and quality considerations and variability over time and space. Basins in which requirements (including environmental requirements) exceed availability and where additional water needs cannot be met without reallocating water from other users, or by improving water-use efficiency, are called closed systems. Open systems, by contrast, are those in which water availability exceeds current requirements and where there is, therefore, still room for expanding water use without environmental damage. Needless to say, management of closed or closing water basins is significantly more challenging than that of open basins. Importantly, integrated management at the basin level does not necessarily imply the need for a basin organization. An analysis by the CA, conducted together with GWP and International Network of Basin Organizations (INBO) (CA, 2008), concluded that ‘‘adaptive, multilevel, collaborative governance arrangements’’ are best able to deal with the complexities of issues that arise at the basin level. Noting that solutions need to be driven by contextual realities and that what works in one basin may not work in another, the CA emphasized that ‘‘Not all water-related problems can or should be solved at the river basin level. Some problems are best addressed at the sub-basin or local level. Others have solutions beyond the basin itself and even outside the water sector, for example in national or federal agricultural policies.’’ Numerous basin-management organizations have been established in different parts of the world. The significance of basin management is reinforced by the INBO, which was established to facilitate exchange of experience and expertise among organizations interested in river-basin management and to promote the principles and means of sound water management. Similarly, there is a vast literature that addresses water management at the water-basin level, particularly at the

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Integrated Water Resources Management

river-basin level. Perhaps the most influential and ambitious publication on the subject is Hooper (2005), which focuses on integrated approaches to river-basin management and develops an integration framework for river-basin management based on principles of natural-resources management and planning. A recent handbook on basin management has been published by GWP and INBO (2009).

There are numerous examples of good practices at the basin level. Boxes 4 and 5 illustrate the different ways in which integrated approaches have been applied at a basin level. Box 4 focuses on the Lerma–Chapala Basin in Mexico, which has faced extreme pressure on its water resources because of many factors, and can therefore be considered a closed system. The Lerma–Chapala Basin has been analyzed

Box 4 The Lerma-Chapala Basin in Mexico. From Hidalgo J and Pen˜a H (2009) Turning water stress into water management success: Experience in the Lerma–Chapala river basin. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development. London and Sterling, VA: Earthscan. Mexico’s Lerma–Chapala basin is one of world’s most over-committed basins, where population growth and industrial and agricultural development have led to a serious imbalance between water availability and water use. It is a classic example of a closed basin, with no natural outflow. The Lerma–Chapala basin is part of the hydrological system formed by the rivers Lerma and Santiago and covers part of the states of Guanajuato, Jalisco, Mexico, Michoaca´n, and Queretaro in central Mexico. Six large cities with more than 1 million inhabitants compete for water in and outside of the basin, and several large industrial corridors lie in the basin and discharge a high concentration of pollutants to the river and water bodies. Over 50% of the land in the basin is used for agriculture, of which a significant amount is irrigated. Water pollution and soil degradation are serious problems, as evidenced by declining soil fertility and increasing soil erosion. The current imbalance in water use began to develop in the 1940s, driven by population growth, heavy construction of water infrastructure, significant industrialization, and large increases in irrigated agriculture. As a result of intensive water use in the middle and lower reaches of the Lerma River, water stopped flowing naturally from Lake Chapala into the Santiago River in the early 1980s. The shrinking volumes of water in Lake Chapala – which dropped 90% in the two decades from 1981 to 2001 – clearly indicate the growth of water demand in the upper basin. While drought undoubtedly contributed to this drop, the main problem appears to be that between the 1940s and 1980s when the federal water authority provided too many water concessions, which led to overexploitation of water sources and a severe water imbalance. Indeed, during the 1990s, less water flowed into Lake Chapala than flowed out through evaporation and water withdrawals, yielding an annual mean deficit of 400 million ha m. To make matters worse, untreated wastewater discharges have degraded the lake’s water quality. To address this situation, the federal government first established the Lerma–Chapala Basin Regional Management, which in the 1980s made efforts to collect more information, improve water plans, define better institutional roles, and involve basin stakeholders in decision making. This led to the first coordination agreement between the federal government and the governments of the five basin states in 1989. A Control and Evaluation Advisory Council (CEAC), a technical working group, designed new rules for surface water reallocation that were approved by consensus by the main stakeholders in 1991. At the same time, a first stage of a water-treatment program was initiated. The CEAC was transformed into the Lerma–Chapala Basin Council (LCBC) in 1993 – the first basin council in Mexico. In 2004, a revised National Water Law strengthened the basin councils, giving them more responsibilities and an organizational structure, and also transformed the river-basin regional management offices into basin organizations, recognizing them as the regional water authorities. At present, the Lerma–Chapala Basin Organization (LCBO) and the LCBC share responsibilities for basin-water governance. Within the context of these institutional changes, a range of instruments have been used over time to tackle the problems caused by water over-use in the basin, including a basin plan and a program of water-resources assessment and use. As a result, Lake Chapala is beginning to recover its natural level, and the Lerma River’s water quality is improving.

Box 5 Basin Planning in the Rio Colorado, Argentina. From Major DC and Lenton RL (1979) Applied Water Resource Systems Planning. Englewood Cliffs: Prentice Hall. The Rı´o Colorado is a relatively small river in southern Argentina that rises near Argentina’s border with Chile, and flows for some 1100 km through a largely arid and sparsely populated area to the Atlantic Ocean, with only minor tributaries. Irrigation is the principal use of water in the basin, although power generation is also important to widely dispersed rural communities. While at present the river waters are used only sparingly, principally for irrigation and hydroelectric power, over the years there have been many proposals for expansion of irrigation and hydroelectric power and for water exports to wine-growing areas north of the basin. The Colorado basin comprises five provinces – Mendoza, Rio Negro, Neuque´n, La Pampa, and Buenos Aires – with different interests in the use of the river’s waters. In the early 1970s, on the initiative of the Argentine government and the five riverine provinces, the Rio Colorado was the site of a pioneering scientific study that aimed to provide a scientific basis for reaching agreement on the allocation of water from the river. The study involved the first complete use of innovative new mathematical modeling techniques for river-basin planning, and indeed the first large-scale application of the approach outlined by the Harvard Water Program (Maass et al., 1962). The study focused on the analysis of alternative development plans, with multiple objectives, with a strong emphasis on the use of mathematical models to analyze such alternatives. As described in Major and Lenton (1979), the study presented a phased set of investment possibilities to provincial and national authorities, including an initial stage of development focused on irrigation, with a reservoir site in the lower basin that has since been developed; a second stage with complementary power; and a third stage with essentially full development of the basin. Importantly, the study formed the basis for a successful 1977 accord on water, energy, and irrigation among the five provinces and the central government of Argentina that has endured for more than 30 years. Since water availability in the Rio Colorado basin (both today and at the time of the study) significantly exceeds requirements, the study’s conclusions and thus the agreement to which it led may not be resilient in the face of possible future changes in water demand and supply. Global changes in agricultural trade, for example, may increase demands for the agricultural and livestock outputs of the Rio Colorado, and increased economic development could lead to increased demand for hydroelectric power. On the supply side, river flows could well decrease as a result of changes in rainfall patterns and the melting of the Andean glaciers. Over time, therefore, increased pressure on the basin’s water resources could force decision makers to consider the prospects for water reallocation and/or water-use efficiency improvements.

Integrated Water Resources Management

extensively, including by Hidalgo and Pen˜a (2009), Dau and Aparicio (2006), Guitro´n et al. (2003), and Wester (2008). Further information can be obtained from Comisio´n Nacional del Agua, 2001. Box 5 focuses on the Rio Colorado Basin in Argentina, an open basin where pressures on the basin’s water resources are fairly low and in which there is still room for meeting additional water needs without reallocating water from existing users.

1.01.6 IWRM at the National Level: Policies and Governance The key feature of water-resources management at the national level is that it focuses not on the physical management of water itself but on the overall governance of the resource and the policies and institutions that facilitate and support management at other levels. Management at the national level therefore deals with strengthening the enabling environment, through measures such as establishing water-resources policies and strategies, improving water legislation, enhancing institutional roles, and strengthening processes for stakeholder participation. The allocation of financial resources (e.g., for investments in water infrastructure) is also clearly a vital task of management at the national level. National-level action can also support the development and use of better instruments for water management, such as those for water-resources assessment. Importantly, national policies and strategies must facilitate and support integrated action at lower levels. While much of what might be considered natural-system integration typically takes place at the basin level, much of what might be called human-system integration takes place at the national level. This includes ensuring that governmental policies take account of water-resource implications and considering water-resource policy within national economic and sectoral policies (GWP Technical Advisory Committee, 2000). Integrating water-resource policy with national economic and sectoral policies, and ensuring that decisions of

15

economic-sector actors are water sensitive, forms another vital set of policymaking at the national level. The need for strengthening water policy as a major issue for sustainable development was among the major accomplishments of the Rio Earth Summit, whose Agenda 21 included a full chapter on freshwater resources and which called for the application of integrated approaches to the development, management, and use of water resources. Attention to national water policies and governance has increased significantly in recent years, and countries the world over have worked to find ways to strengthen their water policies to advance overall sustainable development goals. In Chile, for example, national water policy has played a major role in the country’s overall development since the 1970s. This experience, which has been well documented by Pen˜a et al. (2004), Bauer (2004), GWP Technical Committee (2006), and Pen˜a (2009), is summarized in Box 6. Chile’s example illustrates the importance of aligning water-resource planning and management strategies to overall economic development models. South Africa provides a further example of a country which made a major effort to align its water policies and strategies with national development goals, following its transition to democracy in 1994. This process has been described in Muller (2009), De Coning (2006), and DWAF (2004). In contrast to Chile, however, South Africa attempted to address social, environmental, and economic needs simultaneously, taking advantage of the opportunities to improve water management provided by the country’s broader political-change process (Muller, 2009). There has been a growing literature on the subject of policies and water governance. One landmark publication in this area is Rogers and Hall (2003), who define water governance as ‘‘the range of political, social, economic and administrative systems that are in place to develop and manage water resources, and the delivery of water services, at different levels of society.’’ Recent literature has also examined what might be termed the triggers of water-policy change, and the steps that might be taken to accelerate and sustain positive change. GWP Technical Committee (2009) emphasizes that change is a negotiated

Box 6 National water policy and development: The case of Chile. From Pen˜a H (2009) Taking it one step at a time: Chile’s sequential, adaptive approach to achieving the three Es. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development. London and Sterling, VA: Earthscan. Chile stretches along a narrow strip of land 4200 km long, between the Andes and the Pacific Ocean. The Northern half of the country is arid, while in the Southern half, water is plentiful. Irrigated agriculture accounts for 85% of total consumptive water demand, while domestic uses account for 5%, and mining and industrial uses, around 10% (DGA, 1999). There is significant pressure on existing water resources in the Northern and Central regions of the country, where water requirements are high and water availability is low. At the end of the 1970s, Chile embarked on a new policy of opening up the economy to international trade, promoting the export of products in which the country was competitive. Importantly, nearly all of these export products involved significant water use as part of the production process and were located in areas where water is scarce; for example, copper production was located in the Atacama Desert. As a result, good water-resource management became a prerequisite for the success of the overall export model. Good management in turn required changes in public policies related to water. As described in Pen˜a et al. (2004), Bauer (2004), and GWP Technical Committee (2006), water policy in Chile has evolved considerably since that time, in line with the evolving overall governance structure. Initially, during Chile’s authoritarian period, a water law issued in 1981 with a strong market orientation called for private participation in water supply and sewerage services. A later irrigation law left irrigation initiatives to the private sector, with some partial financing through government subsidies. These steps led to increased efficiency in the use of water in various production processes, as well as increased private investment in Chile’s sewerage and water-supply sector. With the arrival of the democratic period, more attention was given to social, environmental, and regulatory considerations through a range of new laws. In particular, in 2005, the water law was reformed, which called for a more balanced consideration of economic, environmental, and social dimensions and reaffirmed the State’s regulatory role (Pen˜a et al., 2004).

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Integrated Water Resources Management

Table 1

Factors contributing to a conducive environment for change

Particulars

Australia

Chile

Morocco

Namibia

South Africa

Sri Lanka

Water scarcity/conflicts Financial crisis Draughts/salinity Macroeconomic reforms Political reforms Social issues Donor pressures Internal/external agreements Institutional synergy/pressures

**

*

**

**

*

*

*

**

**

***

*

***

***

–

***

*

**

–

***

**

***

–

–

***

–

***

–

***

***

*

*

–

*

**

**

–

–

*

**

*

–

***

***

–

–

*

*

–

**

***

*

*

*

*

The number of *s signifies the relative importance of the factors in the context of each country. ‘‘–’’ means the aspect in question is ‘‘not applicable’’ or ‘‘not evaluated’’. From Saleth RM and Dinar A (2005) Water institutional reforms: Theory and practice. Water Policy 7: 1–19.

political process that is informed by history, public perception, development challenges, and social and economic context; while there are no universally applicable solutions, lessons from experience can provide practical guidance to those involved in change processes. Saleth and Dinar (2005) have identified a number of factors that contribute to a conducive environment for change, as depicted in Table 1. The GWP, the IWMI, and the Stockholm International Water Institute have compiled an extensive set of literature on policy-change processes, the link to which is available in the section titled ‘Relevant websites’. There has also been increasing attention to water in the forums that foster intergovernmental discussions and agreements on sustainable development issues, including in particular the UN Commission on Sustainable Development (UNCSD), established following the Earth Summit in 1992. Recognizing the important role of national policies and strategies in fostering and enabling the development of more integrated management approaches at all levels as called for at the Earth Summit, the UNCSD paid early attention to waterpolicy issues and laid the ground for the World Summit on Sustainable Development (WSSD) in Johannesburg in 2002, which called for all countries to prepare national IWRM and water-efficiency plans by the end of 2005. Subsequently, a significant number of countries embarked on such plans, both in response to the call and as part of national efforts to meet development goals and address specific water and development challenges. Special literature on the subject also emerged at that time, such as Catalyzing Change: A Handbook for Developing Integrated Water Resources Management (IWRM) and Water Efficiency Strategies (GWP Technical Committee, 2004). A growing number of global institutions have focused on fostering better policies and governance at the national level. These include in particular the GWP, which supports countries in the sustainable management of their water resources, and the World Water Council, which promotes international dialog via periodic world water forums and other means. Within the UN system, the United Nations Development Program (UNDP) has given priority to strengthening water governance at the national level, and has established a water-governance facility in Stockholm for this purpose. United Nations Environment Program (UNEP), GWP, and others have played a role in assisting countries to prepare IWRM plans and strategies.

Importantly, the Johannesburg call for the preparation of national IWRM and water-efficiency plans reflected the significant worldwide experience in the development and implementation of national water-resource management policies and strategies. As emphasized by Lenton and Muller (2009), such an experience strongly suggested that national waterresources planning and management must be linked to a country’s overall sustainable development strategy. At the same time, it reinforced the notion that local-level management needs a supportive national-policy framework and support at higher levels.

1.01.7 IWRM at the Transnational and Global Level: Information Sharing, Cooperation, and Technical and Financial Assistance The salient feature of water resources at the transnational and global level is that it focuses on the management of water resources, or a product or resource related to water resources, which crosses national borders. Management at this level thus involves stakeholders in more than one country. The words used to describe transnational and global water management therefore tend to include information sharing, cooperation, negotiation, and technical and financial assistance. Importantly, transnational management complements, rather than replaces, decision making at national and subnational levels. Individual project decisions, for example, are usually taken at national levels, while information gathering and assessment exercises may involve shared efforts. Water resources and products and resources related to water can cross national borders or have an impact in other countries in at least five ways. This translates into at least five different forms of transnational and global water-resources management: 1. Trans-boundary river-basin management. Trans-boundary river-basin management is perhaps the most common form of transnational-level management. It is also the form most often cited in the literature and discussed in international conferences. A very large number of the world’s rivers cross national borders, and many of these involve some degree of trans-boundary management or

Integrated Water Resources Management

2.

3.

4.

5.

cooperation. The Lower Mekong Basin, for example, has witnessed several decades of cooperation in management through the Mekong River Commission and its predecessor, the Mekong Committee (Bruhl and Waters, 2009). Trans-boundary management in the Mekong involves a combination of decision making at different levels. While the Mekong River Commission facilitates information gathering and assessment, individual project decisions are taken by the member countries. The key management challenge here is therefore to manage information sharing, cooperation, negotiation, and conflictresolution activities in ways that increase benefits and/or reduce costs to all parties. Trans-boundary aquifer management. Groundwater resources frequently cross borders in the form of transnational aquifers (see IGRAC, 2009, for a world map of transboundary aquifers). However, despite the large number of transnational aquifers around the world, they have received much less attention in the literature and in international deliberations than trans-boundary river basins (Bourne, 1992). The key management challenge here is generally to ensure the sustainability of resource use by all parties, with assessment and information sharing as major tools. Exports or imports of water resources. Sometimes water resources are exported from one country to another, usually for drinking-water purposes, in pipelines or by other means. Malaysia, for example, exports water resources to Singapore, which in fact depends on imported water for most of its supplies. Since transporting water in bulk over large distances is costly, such imports and exports of water only take place in particular circumstances where other supply options are not feasible or too costly. The key management challenge here usually goes significantly beyond water itself, requiring integration with foreign policy and trade considerations. International trade. Water resources are routinely exported from one country to another via trade in virtual water, which is the term coined to denote the water used in the production of a good or service. In Chile, water-resources management has played a predominant role in the country’s export-oriented economic growth. As noted earlier, the concept of virtual water is much discussed in the literature on trade versus domestic food production. The management challenge here is to ensure that decision making on trade issues takes full account of a country’s water-resource endowments and the impacts that imports or exports of virtual water might have on national water-resource availabilities and requirements. Knowledge management and technical and financial assistance. Managing knowledge and information about waterresources management, as well as technical, institutional, and financial support for water management, generally entails a significant degree of management at regional and global levels. Indeed, much of the focus of the growing number of international forums on water management, such as the UNCSD or the World Water Forums, is on technical assistance, knowledge sharing, capacity building, and financial support. The key challenge here is for countries to coordinate global and regional efforts for

17

knowledge management and technical and financial assistance in ways that best support national policies and strategies to improve water management at all levels. Over the last couple of decades, a number of global institutions have taken on aspects of this challenge. A range of UN organizations assists countries to manage water resources through technical support and capacity building, knowledge sharing, and global monitoring and analysis. The work of these agencies has recently been reinforced by a strengthening of UN-Water, the body which brings together the water activities of all the UN systems and which also oversees the World Water Development Report. Several of the global institutions mentioned earlier are involved in knowledge and information and/or technical, institutional, and financial support for water management, such as the GWP and the World Water Council. The UN Secretary General’s Advisory Board on Water and Sanitation (UNSGAB) helps highlight and mobilize action on global issues that need urgent attention.

1.01.8 IWRM as a Meta-Concept The above-discussed six sections provided an overview of integrated approaches to water-resources management at different levels and helped to unpack the meaning of IWRM at each level. This overview of integrated approaches to waterresources management at different levels suggests that it is best to view IWRM not as a single approach but as a wide range of approaches to manage water and related resources – a metaapproach or meta-concept, as it were, that both transcend the various levels of decision making and recognizes the importance of integrating decision making at each level. Viewing IWRM in this way reinforces the need to understand integration not only in horizontal terms but also in vertical terms, that is, across the different levels of decision making. Clearly, actions at one level should seek to reinforce and complement actions at other levels, within the generally agreed principle that decision making on water resources should be taken at the lowest appropriate level. Importantly, at each level, the approach to integrated management that applies has usually been referred to as an IWRM approach and given a name specifically tailored to that level, such as watershed management or basin management. Some researchers have likewise suggested alternative names to give emphasis to particular attributes, such as adaptive water management (Pahl-Wostl, et al., 2005), which – drawing on concepts from ecosystem management – stresses the need for management approaches that increase adaptive capacity. As we have seen, the way an IWRM approach is given expression differs from level to level. The term ‘integrated’ therefore takes on many forms, depending on the level. Rather than being taken literally as requiring that everything needs to be connected to everything else, the term is best understood as an adjective used to reinforce the notion that intelligent and broad-based management is needed to achieve the goals of sustainable development. In other words, the term integrated is a symbol to describe an approach that goes well beyond integration as such and could be described using other

18

Integrated Water Resources Management

adjectives such as sound, intelligent, broad based or holistic, or systemic. In addition, integration in itself is a multidimensional concept, with the relative importance of each of these dimensions varying from level to level. For example, some aspects of integration within the natural system (e.g., green and blue water, surface and groundwater management) are more relevant at the watershed and basin levels, whereas some aspects of human-system integration (such as considering water-resource policy within national economic and sectoral policies) are more relevant at the national level. Water and development processes also take place at different spatial and temporal scales, which means that integration must also be understood as bridging different spatial and temporal scales, the latter being especially important in the context of climate change. Beyond these differences, however, there are at least four common elements that apply to sound management at all levels and that embody the essence of what an overall IWRM approach is all about: 1. The approach recognizes that water is both a social and an economic good and has multiple uses. 2. The approach seeks to balance multiple objectives that at their core relate invariably to economic efficiency, social equity, and environmental sustainability. 3. The approach entails a broad, holistic, and integrated perspective relevant to the given level of decision making. 4. The approach requires the appropriate involvement of users at the given level of decision making. The body of research and practice to date at all the different levels suggests that there are some elements common to the way in which the approach has evolved at each level as well. For example, while it is now recognized that water-resource management at all levels requires an appropriate combination of hard and soft components – that is, economically, environmentally, and socially sound infrastructure coupled with effective institutions and governance – at most levels, this was not necessarily the starting point. The early stages of watershed management, for example, emphasized engineering approaches to control soil erosion and thus downstream effects, while much of the initial literature on water management at the national level focused on governance and policy issues, at the same time neglecting the need for investments in infrastructural development. Gradually, however, a more balanced approach began to gain prominence at all levels. Similarly, while in the initial stages there was often a tendency toward packages of practices, with a more blueprint approach, in more recent years, a more pragmatic, sequenced, and contextdriven approach has prevailed. At all levels, calls for more integrated forms of management are emphasizing the need to view these as an approach rather than a formal methodology or prescription.

One force identified by Lenton and Muller (2009) was the technical and methodological advances in dealing with complex water systems that began in earnest with the Harvard Water Program carried out at Harvard University from 1955 to 1960. This large multidisciplinary program resulted in the publication of Design of Water Resources Systems: New Techniques for Relating Economic Objectives, Engineering Analysis, and Governmental Planning (Maass et al., 1962). While the program focused principally on developing planning and design methodologies for complex multipurpose water-resource systems at the river-basin level, the approach advanced by the Harvard Water Program combined economic and engineering analysis, drew on techniques of mathematical efficiency models and computer simulation for river systems, and emphasized the role of political processes in decision making. A second force identified by Lenton and Muller (2009) was the recognition of the impact of human activities on the natural environment as crystallized in the Bruntland Report (World Commission on Environment and Development, 1987), and the full expression of these ideas and concepts at the Rio Earth Summit and its far-reaching action document, Agenda 21. Indeed, Chapter 18 of Agenda 21 – the longest chapter in the document – already embodied the concept in its title, which was ‘Protection of the quality and supply of freshwater resources: Application of integrated approaches to the development, management and use of water resources.’ It made a strong case for more integrated approaches than heretofore had been the norm, stating unequivocally that ‘‘the widespread scarcity, gradual destruction and aggravated pollution of freshwater resources in many world regions, along with the progressive encroachment of incompatible activities, demand integrated water resources planning and management.’’ Viewing IWRM as a meta-concept, however, suggests a third and crucial force, which is the set of conceptual and practical advances in management at different levels that have evolved over the last 30 years. Importantly, these advances came not from purely academic centers, but rather from the world of practice and of practical policymaking, reinforced by the findings of field-based research efforts. They have included

•

•

• •

1.01.9 History and Evolution of the Concept of IWRM

• The discussion in the previous section highlights the fact that the concept of an IWRM approach has been an evolving one. Several forces have given impetus to this concept.

Advances in watershed management, led by practical efforts to manage watersheds in a variety of contexts across the globe, the documentation and evaluation of many of these initiatives by organizations such as the World Bank, and the field-based research efforts of institutions such as IWMI and the World Agroforestry Centre. Advances in basin management, led by practical efforts in a number of countries such as France that set up river-basin organizations, field-based research by initiatives such as the CA, and the knowledge-exchange efforts of INBO and others. Advances in AWM, in which the field-based research of IWMI has played a major role. Advances in WSS, fostered by the work of the World Bank’s Water and Sanitation Program and the Water Supply and Sanitation Collaborative Council. Advances in water policy and governance, spearheaded principally by innovative efforts at the national level but supported by organizations such as the GWP and its technical committee.

Integrated Water Resources Management

Importantly, while advances at each level focused on the management issues critical to that level, they all seemed to have some important common elements. In particular, the best practices advocated as appropriate at each level tended to balance multiple objectives – economic efficiency, social equity, and environmental sustainability. They tended to start from a broad, holistic, and integrated perspective relevant to the given level of decision making. In addition, they tended to find a way to involve users at the given level of decision making. Not surprisingly, these common elements are those now recognized as being fundamental to an IWRM approach. While these fairly distinct drivers can be identified, the concept and practice of IWRM evolved over time in a nonlinear way, as is so often the case. While it began with best practices, advances in analytical tools, and advances in specific areas, IWRM as such was formally adopted as a concept at the Rio Earth Summit in 1992.

1.01.10 Assessments and Critiques of the Concept of IWRM As indicated earlier, there is an extensive literature analyzing and critiquing integrated approaches to water-resources management at each of the different levels at which these approaches have evolved in the last several decades – in watershed management, for example, or AWM or basin organizations. In recent years, however, this literature has been supplemented by a significant set of publications on the concept of IWRM as a whole. Initially, much of this literature on IWRM approaches emanated from the GWP, which in 1998 began producing the GWP Technical Committee (TEC) Background Paper series, which address key conceptual issues related to water-resources management (see e.g., Falkenmark, 2003; GWP, 2003, 2008; Rees, 2002, 2006; Rees et al., 2008; and Rogers et al., 1998). Later, the GWP began producing shorter technical and policy briefs on a variety of aspects relating to IWRM, including several that were designed to support countries in their efforts to prepare IWRM and water-efficiency strategies or plans, as well as a set of publications associated with the GWP IWRM ToolBox. Importantly, a range of other authors and groups associated directly or indirectly with GWP also took on the task of explaining and further articulating the concept of IWRM. Examples of these publications include the IWRM tutorial published by CapNet (CapNet, 2008), which provides

19

a brief introductory tutorial of the basic principles of IWRM, and a recent book by Soncini-Sessa et al. (2007), which focuses on the tools available for integrated and participatory water-resources management. Particularly in the last 4 or 5 years, a growing number of researchers and practitioners have taken a harder look at the concept of IWRM, sparking a lively and continuing debate on the subject. Some of these works examine IWRM in a specific geographic context, while others analyze aspects of IWRM and its theoretical underpinnings or critique the IWRM approach and its practical application. Antao and Walkuski (2008, Literature Survey on IWRM, Global Water Partnership, unpublished document) have prepared a comprehensive bibliography on the subject. A comprehensive book edited by Warner (2007) examined multi-stakeholder platforms and their role in integrated management. One important set of publications analyzes aspects of IWRM, including its scientific and theoretical underpinnings. A special issue of the Journal of Contemporary Water Research and Education published in 2006 aimed to provide new insights into the widespread experience of IWRM (Hooper, 2006). In 2007, a special issue of The Geographical Journal focused on what the journal described as Integrated Water Management (IWM); many of the papers in this special issue emphasized that IWM is largely influenced by contextual conditions. Some recent literature (e.g., Pahl-Wostl et al., 2005; Pahl-Wostl and Sendzimir, 2005; and Timmerman et al., 2008) focuses on the concept of adaptive water-resources management, which as explained earlier emphasizes the need for more adaptive systems of management, drawing on ecosystem-management theory. Other recent literature looks at the practical aspects of IWRM. The CA (CA, 2008) notes that in developing countries, what is usually passed-off in the name of IWRM has tended to have a very narrow blueprint package focus – national water policy, a water law and regulatory framework, recognition of the river basin as the unit of planning and management, treating water as an economic good, and participatory management – that represents a big shift from current paradigms and that makes these IWRM initiatives ineffective or counterproductive. Biswas (2004), Biswas and Tortajada (2004), and Biswas (2008) see IWRM as a somewhat fixed approach and raise questions about whether such an approach is practical given the wide range of conditions under which the approach might be applied. While acknowledging some of the above concerns, Lenton and Muller (2009) examine a range of practical examples to

Box 7 Key lessons from IWRM in practice. From GWP Technical Committee (2009) Lessons from Integrated Water Resources Management in Practice, Policy Brief 9. Stockholm: Global Water Partnership. *

* * *

*

IWRM is not a one-size-fits-all prescription and cannot be applied as a checklist of actions. Pragmatic, sensibly sequenced institutional approaches that respond to contextual realities have the greatest chance of working in practice. Water-resource planning and management must be linked to a country’s overall sustainable development strategy and public administration framework. Water management must ensure that the interests of the diverse stakeholders who use and impact water resources are taken into account. Approaches to water-resources management will evolve as the pressures on the resource and social priorities change. The challenge is to support the development of institutions and infrastructure that can meet the challenges of new circumstances. While the river basin is an important and useful spatial scale at which to manage water, there are often circumstances where it is appropriate to work at smaller sub-basin scale or at a regional multi-basin level.

20

Integrated Water Resources Management

see how the principles embodied in the concept of IWRM have been applied at different scales, from very local experiences to reform at the national level and beyond, and give evidence of the positive results obtained. The lessons learned from looking at these examples have been summarized in GWP Technical Committee (2009), and reproduced in Box 7. Taken as a whole, the examples in ‘IWRM in practice’ suggest that, although IWRM has been criticized as a theory that is very difficult to put into practice, it is more like a practice around which it has been very difficult to build a good theory.

Acknowledgments The chapter draws on, and expands, the analytical framework used by the author and his colleague Mike Muller in Integrated Water Resources Management in Practice: Better Water Management for Development (Lenton and Muller, 2009), which greatly facilitated the preparation of this chapter and for which he is very grateful to Mike Muller. The author would also like to acknowledge, with thanks, the very many discussions on the IWRM approach held with colleagues on the Technical Committee of the Global Water Partnership from 2003 to 2006. These discussions, as well as the many background papers and briefs emanating from the Technical Committee during this period, have substantially contributed to the conceptual thinking and information contained in this chapter.

References Agarwal A and Narain S (1999) Community and household water management: The key to environmental regeneration and poverty alleviation. Presented at EU UNDP Conference. Brussels, February 1999. Barry B, Namara R, and Bahri A (2009) Better rural livelihoods through improved irrigation management: Office du Niger (Mali). In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development, pp. 71–87. London and Sterling, VA: Earthscan. Bauer C (2004) Siren Song: Chilean Water Law as a Model for International Reform. Washington, DC: Resources for the Future. Biswas AK (2004) Integrated water resources management: A reassessment. Water International 29: 248--256. Biswas AK (2008) Integrated water resources: Is it working? Water Resources Development 24(1): 5--22. Biswas AK and Tortajada C (eds.) (2004) Appraising the Concept of Sustainable Development: Water Management and Related Environmental Challenges. Oxford: Oxford University Press. Bourne C (1992) The International Law Commission’s draft articles on the law of international watercourses: Principles and planned measures. Colorado Journal of International Environmental Law and Policy 3: 65--92. Bruehl H and Waters M (2009) Transboundary cooperation in action for integrated water resources management and development in the Lower Mekong Basin. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development, pp. 189–204. London and Sterling, VA: Earthscan. CA (Comprehensive Assessment of Water Management in Agriculture) (2007) Water for Food, Water for Life: A Comprehensive Assessment of Water Management in Agriculture. London: Earthscan; Colombo: International Water Management Institute. CA (Comprehensive Assessment of Water Management in Agriculture) (2008) Developing and managing river basins: The need for adaptive, multilevel, collaborative institutional arrangements, Issue Brief No. 12, International Water Management Institute and Global Water Partnership. CapNet (2008) IWRM Tutorial. http://www.archive.cap-net.org/iwrm_tutorial/ mainmenu.htm (accessed March 2010).

Comisio´n Nacional del Agua (2001) Programa Hı´drico de la cuenca Lerma-Santiago 2001–2006. Me´xico. CSE (Centre for Science and Environment) (1994) Partners in prosperity. Down to Earth, 15 February 1994. CSE (Centre for Science and Environment) (1998) Sukhomajri at the crossroads. Down to Earth, 15 December 1998. CSE (Centre for Science and Environment) (2002) Foisting failure Down to Earth, 31 August 2002. CSE (Centre for Science and Environment) (2007) Saga of two villages Down to Earth, 15 November 2007. Darghouth S, Ward C, Gambarelli G, Styger E, and Roux J (2008) Watershed Management Approaches, Policies, and Operations: Lessons for Scaling Up, Water Sector Board Discussion Paper Series, Paper No. 11. Washington, DC: The World Bank. Dau FE and Aparicio MJ (eds.) (2006) Acciones para la recuperacio´n ambiental de la cuenca Lerma-Chapala. In: Comisio´n Estatal de Agua y Saneamiento, Gobierno del Estado de Jalisco, 126 pp. Me´xico: Comisio´n Estatal de Agua y Saneamiento de Jalisco, Guadalajara. De Coning C (2006) Overview of the water policy process in South Africa. Water Policy 8: 505--528. de Fraiture C, Giordano M, and Liao Y (2008) Biofuels and implications for agricultural water use: Blue impacts of green energy. Water Policy 10(supplement 1): 67--81. Direccio´n General de Aguas (DGA) (1999) Repu´blica de Chile, Polı´tica Nacional de Recursos Hı´dricos. DWAF (2004) National Water Resource Strategy, 1st edn. Pretoria: Department of Water Affairs and Forestry. Falkenmark M (2003) Water Management and Ecosystems: Living with Change, TEC Background Papers No. 9. Stockholm: Global Water Partnership. Farrington J, Turton C, and James AJ (1999) Participatory Watershed Development: Challenges for the Twenty-First Century. New Delhi, India: Oxford University Press. Guitro´n A, Hidalgo J, Aparicio J, and Aldama A´ (2003) A water crisis management: The Lerma-Chapala basin case. In: Brebbia CA (ed.) Water Resources Management II, pp. 345--354. Southampton: WIT Press. GWP (Global Water Partnership) (2003) Integrated Water Resources Management Toolbox, Version 2. Stockholm: GWP Secretariat. GWP (Global Water Partnership) and INBO (International Network of Basin Organizations) (2009) A Handbook for Integrated Water Resource Management in Basins. GWP Technical Advisory Committee (2000) Integrated Water Resources Management, TAC Background Papers No. 4. Stockholm: Global Water Partnership. GWP Technical Committee (2004) Catalyzing Change: A Handbook for Developing Integrated Water Resources Management (IWRM) and Water Efficiency Strategies. Stockholm: Global Water Partnership. GWP Technical Committee (2006) Water and Sustainable Development: Lessons from Chile, Catalyzing Change Series Policy Brief 2. Stockholm: Global Water Partnership. GWP Technical Committee (2009) Lessons from Integrated Water Resources Management in Practice, Policy Brief 9. Stockholm: Global Water Partnership. Heathcote IW (1998) Integrated Watershed Management: Principles and Practices. p. 441. New York, NY: Wiley. Hellegers P, Zilberman D, Steduto P, and McCornick P (2008) Interactions between water, energy, food and environment: Evolving perspectives and policy issues. Water Policy 10(supplement 1): 1--10. Hidalgo J and Pen˜a H (2009) Turning water stress into water management success: Experience in the Lerma–Chapala river basin. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development, pp. 107–120. London and Sterling, VA: Earthscan. Hooper B (2005) Integrated River Basin Governance: Learning from International Experiences. London: IWA Publishing. Hooper B (2006) Integrated water resources management: Governance, best practice, and research challenges. Journal of Contemporary Water Research and Education 35: 1--7. IGRAC (International Groundwater Resources Assessment Centre) (2009) Transboundary Aquifers of the World Map – Update 2009. Joshi PK, Jha AK, Wani SP, Joshi L, and Shiyani RL (2005) Meta-analysis to assess impact of watershed programme and people’s participation. In: Comprehensive Assessment of Water Management in Agriculture, Research Report 8. Colombo: International Water Management Institute. Kerr J (2002) Sharing the benefits of watershed management in Sukhomajri, India. In: Pagiola S (ed.) Selling Forest Environmental Services: Market-Based Mechanisms for Conservation and Development. London: Earthscan. Khurana MR (2005) Common property resources, people’s participation and sustainable development: A study of Sukhomajri. Panjab University Research Journal (Arts) XXXII(18.2). April–October.

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Lenton R and Muller M (2009) Integrated Water Resources Management in Practice: Better Water Management for Development. London and Sterling, VA: Earthscan. Lenton R and Walkuski C (2009) A watershed in watershed management: The Sukhomajri experience. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development. London and Sterling, VA: Earthscan. Lenton R, Wright A, and Lewis K (2005) Health, Dignity and Development: What Will It Take?, pp. 17–28. London: Earthscan. Maass A, Hufschmidt MM, Dorfman R, Thomas HA Jr, Marglin SA, and Fair GM (1962) Design of Water-Resource Systems; New Techniques for Relating Economic Objectives, Engineering Analysis, and Governmental Planning. Cambridge: Harvard University Press. Major DC and Lenton RL (1979) Applied Water Resource Systems Planning. Englewood Cliffs: Prentice Hall. Muller M (2009) Attempting to do it all: How a New South Africa has harnessed water to address its development challenges. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development, pp. 169–185. London and Sterling, VA: Earthscan. Pahl-Wostl C, Downing T, Kabat P, et al. (2005) Transition to Adaptive Water Management: The NeWater Project. NeWater Working Paper 1, Institute of Environmental Systems Research, University of Osnabru¨ck. Pahl-Wostl C and Sendzimir J (2005) The Relationship between IWRM and Adaptive Water Management. NeWater Working Paper 3. Institute of Environmental Systems Research, University of Osnabru¨ck. Pen˜a H (2009) Taking it one step at a time: Chile’s sequential, adaptive approach to achieving the three Es. In: Lenton R and Muller M (eds.) Integrated Water Resources Management in Practice: Better Water Management for Development, pp. 153–168. London and Sterling, VA: Earthscan. Pen˜a H, Luraschi M, and Valenzuela S (2004) Water, Development and Public Policies. South American Technical Advisory Committee (SAMTAC), Economic Commission for Latin America and the Caribbean (ECLAC) and Global Water Partnership (GWP). Rees JA (2002) Risk and Integrated Water Management, TEC (formerly TAC) Background Papers No. 6. Stockholm: Global Water Partnership. Rees JA (2006) Urban Water and Sanitation Services: An IWRM Approach, TEC Background Papers No. 11. Stockholm: Global Water Partnership. Rees JA, Winpenny J, and Hall AW (2008) Water Financing and Governance, TEC Background Papers No. 12. Stockholm: Global Water Partnership. Rogers P, Bhatia R, and Huber A (1998) Water as a Social and Economic Good: How to Put the Principle into Practice, TAC Background Papers No. 2. Stockholm: Global Water Partnership. Rogers P and Hall A (2003) Effective Water Governance, TEC Background Papers No. 7. Stockholm: Global Water Partnership. Rosegrant M, Cai X, and Cline S (2002) World Water and Food to 2025: Dealing with Scarcity. Washington, DC: IFPRI. Saleth RM and Dinar A (2005) Water institutional reforms: Theory and practice. Water Policy 7: 1--19. Seckler D (1986) Institutionalism and agricultural development in India. Journal of Economic Issues XX(4). December.

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Shah T (2009) Taming the Anarchy: Groundwater Governance in South Asia. Washington, DC: Resources for the Future. Colombo: International Water Management Institute. Sharma BR, Samra JS, Scott CA, and Wani SP (eds.) (2005) Watershed Management Challenges: Improving Productivity, Resources, and Livelihoods. New Delhi, India: International Water Management Institute. Soncini-Sessa R, Cellina F, Pianosi F, and Weber E (2007) Integrated and Participatory Water Resources Management – Practice, Volume 1b. Developments in Integrated Environmental Assessment Series. Amsterdam: Elsevier. Timmerman JG, Pahl-Wostl C, and Moltgen J (2008) The Adaptiveness of IWRM: Analysing European IWRM Research. London: IWA Publishing. UNCED (United Nations Conference on Environment and Development) (1992) Agenda 21, Report of the United Nations Conference on Environment and Development. http://www.un.org/esa/sustdev/documents/agenda21 (accessed March 2010). Warner J (ed.) (2007) Multi-Stakeholder Platforms for Integrated Water Management Studies in Environmental Policy and Practice. Aldershot, UK: Ashgate Publishing. Wester P (2008) Shedding the Waters: Institutional Change and Water Control in the Lerma-Chapala Basin, Mexico. PhD Dissertation, Wageningen University, Wageningen, The Netherlands. World Bank (2003) Tunisia Northwest Mountain Areas Development Project Performance Assessment Report. Washington, DC: World Bank. World Bank (2004a) Peru Sierra Project Implementation Completion and Results Report. Washington, DC: World Bank. World Bank (2004b) Tajikistan Community Agriculture and Watershed Management Project, Project Appraisal Document. Washington, DC: World Bank. World Bank (2004c) Turkey Eastern Anatolia Project, Project Performance Assessment Report. Washington, DC: World Bank. World Bank (2005) China Loess II Project Implementation Completion and Results Report. Washington, DC: World Bank. World Bank (2006) Water Management in Agriculture: Ten Years of World Bank Assistance, 1994–2004. Washington, DC: Independent Evaluation Group (IEG)/The World Bank. World Commission on Environment and Development (1987) Our Common Future (Brundtland Report). Oxford: Oxford University Press. WSSD (World Summit on Sustainable Development) (2002) Johannesburg Plan of Implementation. http://www.un.org/esa/sustdev/documents/WSSD_POI_PD/ English/POIToc.htm (accessed March 2010).

Relevant Websites http://www.rlc.fao.org Food and Agricultural Organization of the United Nations; Regional Office for Latin America and the Caribbean; Network on Watersheds Management. http://www.gwpforum.org Global Water Partnership. http://rupes.worldagroforestry.org Rewards for, Use of and shared investment in Pro-poor Environmental Services.

1.02 Governing Water: Institutions, Property Rights, and Sustainability E Schlager and C Bauer, The University of Arizona, Tucson, AZ, USA & 2011 Elsevier B.V. All rights reserved.

1.02.1 1.02.2 1.02.3 1.02.3.1 1.02.3.2 1.02.4 1.02.5 References

Introduction International Organizations and Water Policy Debate Governing Water from the Ground Up Local Communities, Property Rights, and Water Linking Water Uses and Administration across Multiple Scales and Jurisdictions Courts: Hiding in Plain View Conclusion: Reconceptualizing Water Governance

1.02.1 Introduction What is water governance and why is it included in this treatise? Water governance is one of those elastic terms that has something for everyone – everyone agrees that it is important and many disagree about what it means (‘institutions’ is another example). In a general way, governance refers to how people make decisions and govern themselves, whether in organizations or at the larger scale of societies. Thus, governance is about social, political, and economic processes and how they interact over time. This of course includes the realm of formal governmental institutions; however, it goes beyond this realm as well to include other aspects of social and political life. Is governance any different from politics? It depends on what one thinks politics means. They are not easily distinguishable, especially since both terms tend to be used with a great deal of abstraction and generality. In much international debate about water policy, water governance has become a black box, where we put important but complicated issues that we are not sure how to think about or talk about. Some of these issues are so deeply political that they are dangerous to discuss in public, and governance is soothing and blurs the edges. Our goal in this chapter is to open the lid of the black box of water governance and look more closely at what is inside. Where is the basic political economy, the distribution of wealth, and power in society? Who gains and who loses in water governance? What are people talking about when they talk about governance? What analytical tools can we bring to bear that would allow us to flesh out and make sense of governance?

1.02.2 International Organizations and Water Policy Debate We begin by reviewing the positions on water governance of several prominent international organizations. A good place to start is Peter Rogers and Alan Hall’s paper on water governance for the Global Water Partnership, or GWP (Rogers and Hall, 2003; building on Rogers’ work for the Inter-American Development Bank (Rogers, 2002)). The GWP is an international organization that was established in 1996 to promote integrated water resources management (IWRM) around the

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world. IWRM is an international catch phrase that refers to the idea that water should be managed in a holistic, comprehensive, and multidisciplinary way – a way that does justice to the hydrologic cycle. This usually means focusing on the relationships between water quality and quantity, between surface water and groundwater, integrating across different water-using sectors at the level of river basins. Since the GWP aims to be the official, mainstream voice of IWRM around the world, its position on water governance is worth looking at (Bauer, 2004; Conca, 2006; GWP, 2000b). According to the GWP, ‘‘IWRM is a process which promotes the coordinated development and management of water, land and related resources, in order to maximize the resultant economic and social welfare in a equitable manner without compromising the sustainability of vital ecosystems’’ (GWP, 2000a: 22). Rogers and Hall begin by underlining the intensely political nature of allocating water, and then state that ‘‘governance is about effectively implementing politically achieved allocations’’ (Rogers and Hall, 2003: 4). Governance also ‘‘broadly embraces the formal and informal institutions by which authority is exercised’’ (Rogers and Hall, 2003: 7). They quote contemporary definitions of governance by the United Nations Development Program and the GWP that are breathtaking in their sweep: Governance is the exercise of economic, political and administrative authority to manage a country’s affairs at all levelsyit comprises the mechanisms, processes and institutions through which citizens and groups articulate their interests, exercise their legal rights, meet their obligations and mediate their differences. (UNDP, 2001)

Water governance refers to the range of political, social, economic, and administrative systems that are in place to develop and manage water resources, and the delivery of water services, at different levels of society. (GWP, 2002)

On such a wide-open playing field, Rogers and Hall run through a series of major issues, approaches, and schools of thought in political science and political and social theory. Examples include the incentives for legislators’ behavior, the relationship between the state and civil society, and the tension between markets and centralized hierarchies. Where they focus on water governance in particular, they emphasize the legal foundations and different forms of property rights and

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Governing Water: Institutions, Property Rights, and Sustainability

institutional arrangements. They argue that water governance is affected by factors both internal and external to the water sector itself. Finally, they make the link from water governance to the need for water policy reforms and IWRM, along the lines expressed by the GWP mission. The GWP has published other documents that have shaped the international water policy debate. In its framing paper for the 2nd World Water Forum in 2000, the GWP famously declared ‘‘The water crisis is mainly a crisis of governance.’’ Governance here means conflict resolution: ‘‘The present threat to water security lies in the failure of societies to respond to the challenge of reconciling the various needs for and uses of watery. And governance lies at the center of the tension and delicate balance between different water uses and their management’’ (GWP (2000b: 23), emphasis in original; see also Cosgrove and Rijsberman (2000)). From this angle, that is, to deal with conflict, the GWP argues for a series of policy reforms to improve water governance. One key reform is to put IWRM into practice, although here too the GWP underlines the political nature of ‘‘integrating the seemingly incompatible goals, beliefs, interests, and knowledge of every water user’’ (GWP, 2000b: 25). Another key reform is to strengthen institutions and management, which will clearly require the organization and exercise of political power. Two other reforms to improve governance are to promote transparency and participation and to take a more economic approach to valuing water and designing price incentives. In short, the GWP position on water governance revolves around recognizing political conflicts and developing the capacity – that is, power – to handle them. As befits an international organization, the GWP documents are silent or vague about any specific political alliance or opposition in a given country (GWP, 2000a). In the year 2000, the other landmark event in global water governance (besides the 2nd World Water Forum) was the final report of the World Commission on Dams, or WCD (World Commission on Dams, 2000). The WCD was an unusual organization in many ways, a pioneering example of building multi-stakeholder and international consensus about complex and conflictive issues. The story and process of the WCD are as significant as the conclusions. The WCD was created through negotiations between the World Bank, several national governments, and a group of international environmental nongovernmental organizations (NGOs), and its task was to evaluate the world experience of large dams. The commissioners, once named, had to hire staff and decide how to carry out their tasks, including how to build legitimacy and authority for their results and recommendations. The Commission itself had no regulatory power and it was dissolved after disseminating its final report (the details are well delineated in Dubash et al. (2001) and Conca (2006)). The WCD report looked at water governance as a matter of social equity and justice above all: conflict is understood as driven by who wins and who loses. The report’s broad scope is announced in the second sentence: ‘‘The debate about dams is a debate about the very meaning, purpose and pathways for achieving development.’’ The WCD proposes a new framework for decision making, describing an approach based on recognizing different rights and assessing different risks involved with dams. This rights-and-risks approach aimed to broaden

the range of who counted as legitimate stakeholders in negotiating problems about dams, as well as broadening the issues on the agenda. The WCD argument is in part a moral argument. The Commission’s concrete impacts remain uncertain. Let us conclude this brief review with the World Bank. In the World Bank’s last two major documents about water policy, in 1993 and 2004, the authors use the term management instead of governance, but they are writing about the same issues. Both documents use the language of contemporary IWRM, while adding or strengthening the promarket stamp that has come with the World Bank in recent decades (e.g., World Bank (2004) adopts the GWP image of the comb to illustrate IWRM: the teeth of the comb represent different water-using sectors and the handle holds them together). However, the World Bank’s position on water politics and political economy has changed since the early 1990s, with significant implications for water governance (even if the World Bank calls it management). The World Bank’s 1993 Water Policy Paper had something for almost everyone: a great degree of emphasis on markets, privatization, and pricing, coupled with arguments in favor of strong government regulation and strengthened institutional arrangements (World Bank, 1993: 40). The paper spoke in grand generalities about IWRM and a so-called comprehensive analytical framework for resolving water problems, but the concrete meaning was vague and the political tone was muted. A decade later, in contrast, the World Bank’s key water experts were reasserting their authority and perspective after years of defending the World Bank from outside criticism. The 2004 Water Resources Sector Strategy focused on the implementation of the World Bank’s ideas and policies, which means a more down-to-earth and pragmatic approach to how reforms play out in the real world. The key is recognizing that water resources management is intensely political and that reform requires the articulation of prioritized, sequenced, practical and patient interventions. To be a more effective partner, the Bank must be prepared to back reformers and to pay more explicit attention in design and implementation to the political economy of reform. (World Bank, 2004: 3, emphasis added)

What does that mean? Later in the document the ‘‘political economy of water management and reform’’ is described as placing ‘‘particular emphasis on the distribution of benefits and costs and on the incentives that encourage or constrain more productive and sustainable resource use.’’ (World Bank, 2004: 13, emphasis added) This bare-knuckle approach means that ‘‘the World Bank will re-engage with high reward/high risk hydraulic infrastructure,’’ that is, dams (World Bank, 2004: 3). The World Bank’s maneuver is a notable dodge: having invoked political economy as the distribution of benefits and costs, the World Bank reframes the debate in terms of risks and rewards and then is silent about how the risks and rewards are distributed. If we want to know, we will have to find out for ourselves. The challenge for researchers is to identify specific reformers whom the World Bank has backed, and specific examples of water reforms, and then analyze the distribution of benefits and costs: Who gained and who lost? Also, how was that related to the reforms’ design and implementation?

Governing Water: Institutions, Property Rights, and Sustainability

The key international actors have recognized the centrality of politics in water governance, but have avoided tackling the more difficult issues of who participates in making collective decisions, the types of authority participants have to address problems, issues, and conflicts, and how benefits and burdens are distributed among people. While IWRM is often pointed to as the model for governance, it is also largely content free (Conca, 2006). Furthermore, IWRM possesses a distinct topdown bias (Kemper et al., 2007; Carlsson and Berkes, 2005). It assumes that a central government shares its power to make and enforce decisions with other, lower-level governments, and that civil society – in whatever form that takes: water user associations, nonprofit organizations, etc. – is invited into the decision-making processes. In the following section, we propose an alternative starting point, water users, and a form of policy analysis that attempts to diagnose the politics of specific water settings before prescribing institutional reforms.

1.02.3 Governing Water from the Ground Up 1.02.3.1 Local Communities, Property Rights, and Water A major research program in the social sciences focuses on the study of common pool resources, such as rivers, streams, and groundwater basins and the riparian and aquatic habitat such water sources support (Blomquist, 1992; Ostrom et al., 2002). One of the defining lessons from more than two decades of this research is the centrality of local-level resource users. Sustainable use of resources requires active participation of local water users in water governance. Fundamentally, common pool resources present two challenges to resource users and others engaged in management and governance – exclusion and use (Ostrom et al., 1994). Since it is challenging to realize and sustain exclusion, and since subtractability means resource units harvested by one user are not available for another user, common pool resources tie users together. Realizing exclusion and limiting the number of resource units subtracted from a common pool resource require the cooperation of most resource users. The benefits and costs resource users achieve in using a common resource depend on the actions of all. If sustainable use of common pool resources is to be realized, then both exclusion and subtractability must be carefully considered and managed. For instance, a common approach to address declining groundwater tables among states in the western US is to adopt well moratoria. Such a policy tool directly addresses exclusion: only those resource users with wells are allowed to access groundwater basins and no new resource users are allowed in. This policy, however, does not address problems raised by subtractability. If a well moratorium is not matched with limits on the amount of water each well may pump, a groundwater basin may be mined by existing users. Conversely, placing limits on how much water may be withdrawn from a groundwater basin, and by what means, without addressing exclusion, also exposes the basin to mining. For instance, the Arizona Groundwater Management Act, as originally written, limited water use by municipal water utilities to a specified number of gallons per person per day.

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While gallons per person per day mandates directly addressed subtractability issues, they did not address exclusion issues. In effect, any person could move to a municipality governed by the groundwater act and through the municipal water provider access the groundwater basin. No limits were placed on the number of people that water utilities could serve, thus allowing access to anyone to the groundwater basin. Until relatively recently, a common belief among policy analysts and policymakers was that local-level resource users could not adequately address exclusion and subtractability issues, that is, they could not sustainably govern their shared water resources. Rather, the expectation was that most common pool resource settings were characterized by the rule of capture and, consequently, a race to harvest as many resource units as possible before they were captured by others (Hardin, 1968; Olson, 1965; Ostrom, 1990). The result was over-harvesting at best, and severe degradation or destruction of a common pool resource at worst, and the need for external intervention to save resource users from the race to harvest. It turns out, however, that the models predicting overuse and degradation were too simple, failing to adequately capture key features of many common pool resource settings, such as the ability of resource users to communicate with one another, to share experiences and knowledge of the resource; or the capacity of resource users to develop norms of reciprocity and sharing; or the values that resource users place on the resource and on the opportunity to have their children and grandchildren use the resource; or the experience resource users have in governing other areas of their lives that they can transfer to governing a common pool resource, and so on and so forth (Ostrom, 2007). For the past two decades, hundreds of cases of resource users sustainably managing water resources have been documented and published (Digital Library of the Commons, 2009). In a number of instances, local-level resource users have developed governing arrangements that perform better than government regulation and management. For instance, in an early study comparing farmer-managed irrigation systems and government-managed irrigation systems, Tang (1994) found that among irrigation systems that performed well, rules that govern water allocation and maintenance activities are better crafted to the specific conditions of each irrigation system. High-performing systems, which were more likely to be farmer managed, were associated with multiple rules that adequately limited access to the system and that fairly allocated water among the irrigators. In other words, farmers paid careful attention to exclusion and subtractability. Poorly performing irrigation systems, which were more likely to be government managed, were characterized by a single simple rule set or by no rules at all. Access to the irrigation systems was not adequately regulated and water allocation rules often did not work well. Monitoring and enforcement systems also differ between irrigator-owned systems and government-owned systems. Government-owned systems relied on full-time, paid guards. Farmer-owned systems relied on unpaid part-time guards (Tang, 1994: 241). However, guards in farmer-owned systems were much more likely to impose sanctions on rule breakers than were guards in government-owned systems. Furthermore, rule-following behavior was much more common in

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Governing Water: Institutions, Property Rights, and Sustainability

farmer-owned systems than in government-owned systems, whether guards were present or not (Tang, 1994: 241). Farmers who participate in devising their own irrigation rules are much more likely to follow and actively monitor and enforce their rules. The means by which water users address exclusion and use issues is through devising, following, monitoring and enforcing institutional arrangements. By institutional arrangements we refer to strategies, norms, rules, and property rights systems (Ostrom et al., 1994; Ostrom, 2007). Property rights define relationships among people in relation to things, such as a common pool resource and the resource units it produces. For every right an individual holds, rules authorize or forbid specific actions in exercising the right (Schlager and Ostrom, 1992: 250). For instance, among the irrigators studied by Tang (1992, 1994), those irrigators in farmer-managed systems exercised their rights of access to and withdrawal of water from the irrigation system in substantially different ways than irrigators who were part of a government-managed system. Access rights to government-managed systems were operationalized by a single rule – ownership of land in the irrigation command area (Tang, 1994: 231). Access rights to farmer-managed systems were operationalized by a variety of rules, such as purchasing or leasing shares in the irrigation system, becoming a member in an irrigation organization, or paying fees for access to water (Tang, 1994: 231). Similarly, withdrawal rights in government-managed systems were also operationalized by a single rule, whereas farmer-managed systems were characterized by multiple rules. Furthermore, irrigators in farmer-managed systems also exercised rights of management and exclusion (Schlager and Ostrom, 1992: 251), that is, irrigators possessed the right to regulate how water would be allocated and how maintenance activities would take place (management rights), as well as to determinine who would hold rights of access and how such rights would be exercised (exclusion rights). Irrigators in government-managed systems did not exercise rights of management and exclusion; rather those rights were exercised by government officials. Rights of management and exclusion are collective-choicelevel rights (Schlager and Ostrom, 1992). Holders of such rights are authorized to develop rules that define how rights of access and withdrawal may be exercised. Who holds rights of management and exclusion and their relationship to the common pool resource affect the types of rules devised. Government officials, even as they exercise rights of management and exclusion, will not exercise rights of access and withdrawal. Since they are not directly subject to the irrigation rules they devise, they face few incentives to design rules that ensure the effective operation of irrigation systems. Instead, they may devise rules for other purposes, for instance, to increase their political support or lighten their administrative burdens. Conversely, because farmers in farmer-managed irrigation systems directly experience the consequences of their rule-making decisions, they confront incentives to craft the rules to the particular situation that they face (Tang, 1992). What common pool resource studies have demonstrated is the diversity and the multi-dimensionality of institutional arrangements devised by water users. In any given common pool resource setting, resource users are likely to hold different

bundles of property rights and exercise them in different ways, depending on the rules that specify different actions. In other words, legal pluralism, that is, multiple property rights systems, is likely to be the rule and not the exception across water settings. For instance, irrigation ditches and districts in eastern Colorado often control a portfolio of water that consists of several different types and sources, each governed by a different set of property rights and rules. Water from streams and rivers is governed differently from federal project water, and federal project water is governed differently from irrigation district reservoir water (Blomquist et al., 2004). Rivers and streams are governed by the prior appropriation doctrine, federal project water by the project’s enabling legislation, and district reservoirs by the rights and rules developed by the districts’ governing boards. In addition, the institutional arrangements devised by water users are multidimensional. Resource users give consideration to the transactions costs of implementing and administering property rights and rules, the allocation of benefits and costs realized by the rights and rules, how the rights and rules allocate risk, and whether the rights and rules are likely to dampen conflict among resource users. In other words, whether rules are effective is evaluated along multiple dimensions. A classic illustration of the multidimensionality of institutional arrangements involves the irrigation systems built and managed by farmers, located in Ilocos Norte, the Philippines, as reported by Coward (1979). Farmers gain entry into the systems by purchasing shares. A share entitles a household to a rich set of property rights – access, withdrawal, exclusion, and management. The rules the farmers have devised for exercising their property rights and engaging in irrigated farming are sophisticated. A share entitles a household to several plots of land dispersed along a canal so that all farmers have land located closer to the more desirable head of the canal and land located closer to the less desirable tail end of the canal. Not only do such rules dampen conflict that often emerges among farmers in different locations of an irrigation system, but they also allow the farmers to spread the risk of water shortages. During extremely dry periods, a portion of the irrigation system may be shut down, and farmers forbidden from irrigating plots located in the closed sections. However, farmers still have use of plots located in open sections. The shares allocate risks, benefits, and costs in a proportionate manner. Water and work obligations for maintaining the systems are allocated based on the proportion of land encompassed by shares. More land translates into not only more water but also greater work obligations. Finally, attention is paid to monitoring of water use. Water monitors are selected from among farmers and they are paid for their services by grants of land at the tail end of canals. Whether those plots receive water depends in part on how well the irrigators on the canal are monitored. The Filipino irrigation systems are just one example of many by which resource users have devised institutional arrangements that are self-reinforcing, that is, the decisions and trade-offs made along one dimension (e.g., allocating plots of land across an irrigation command area), supporting and reinforcing decisions made along another dimension (e.g., allocating the risk of water shortages), and so on. Many other such irrigation examples may be pointed to, such as Bali

Governing Water: Institutions, Property Rights, and Sustainability

(Lansing, 1991), Nepal (Lam, 1998), and Spain (Maass and Anderson, 1986). Cases such as these illustrate the institutional artisanship that resource users are capable of. As Ostrom (1990, 1999, 2007) has repeatedly noted, there is no single best set of property rights nor single rule set that support sustainable uses of water. Rather, long-enduring institutional arrangements appear to share some common features (Ostrom, 1990). According to Ostrom (1990), the most important feature is exclusion of nonowners. Exclusion is critical if water users are to commit to following a set of institutional arrangements over time and investing in modifying them as circumstances warrant. Water users must be assured that they will capture the benefits of their actions. Exclusion, however, while critical, is insufficient to ensure long-term commitment to property rights and rules. The institutional arrangements must be appropriate, crafted to the exigencies of the situation, and as the situation changes, the resource users must have the ability to modify the rules. Accountable monitors and graduated sanctioning maintain water users’ commitment to institutional arrangements. Finally, conflict-resolution mechanisms and at least a minimal recognition of the right to organize prevent these institutional arrangements from unraveling due to internal strife or invasion from external governmental authorities (Ostrom, 2007). While it is clear that water users are capable of designing relatively sophisticated and resilient institutional arrangements, there are just as many instances of water users failing to develop or sustain governing arrangements for many reasons. Resource users may not have the capacity, experience, or incentives to overcome collective action problems and devise governing arrangements that allow them to sustainably use their water resources (see Ostrom, 1999, 2000, 2007). Alternatively, certain types of resources or resource problems are extremely difficult for local resource users to address (Schlager and Blomquist, 2005; Schlager, 2005). For instance, groundwater basins present extraordinarily difficult challenges that resource users struggle with. Boundaries of basins are not easily identified, if at all, by resource users, neither is the structure of the resources. Groundwater pumpers, even if they are in close proximity of one another, may not know whether they are pumping from the same basin, or, if they are, whether they are pumping from the same aquifer within the basin. They are also unlikely to know whether and to what extent their basin is connected to other basins, or whether and in what ways their basin is connected to surface water sources. In other words, groundwater pumpers cannot easily sketch out the boundaries and dimensions of the resource or the boundaries of the resource users (Schlager and Blomquist, 2005). In addition, a groundwater basin, even if its boundaries and basic structure are well specified, is not just a bathtub filled with water. It consists of different pieces, such as layers of aquifers, connected surface water sources, recharge areas, flow paths and impediments, and so forth, structured, connected, and disconnected by varying geological formations. The pieces are characterized by different temporal and spatial scales. For instance, depending on soil characteristics, water percolating from the land surface will reach different portions of an aquifer at differing rates. Some water quality or water supply

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impacts will be felt quickly and others gradually, some throughout an aquifer and others in localized zones. The uncertainties around boundaries of the groundwater basins and the identities of groundwater pumpers combined with the complexity of the resource create substantial barriers for groundwater users to organize and develop institutional arrangements that adequately address exclusion and subtractability issues. Encouraging resource users to collectively limit their harvesting activities will be difficult if they are not assured that they will reap the benefits of conservation (Schlager and Blomquist, 2005). Finally, even if resource users do develop governing arrangements, those arrangements may rest on values or center on goals in conflict with those of the society at large. Segments of communities may be excluded from a water resource or from participating in decision-making processes based on ethnicity, gender, or class (Ilahiane, 1999). Allocation rules may favor one group of users over another (Tang, 1992), and enforcement of rules may rest on questionable practices (Rose, 2002). Each of these issues, a lack of local self-governance, the inability to tackle particular types of problems, and the pursuit of questionable values or goals, draws attention to the larger context in which local governance operates. What types of resources, authorities, and institutional ties link resource users with governments and organizations at the regional, national, or international levels that may be drawn upon to support investment in local governance capacity, or assist local resource users in addressing regional problems, or allow higherlevel officials to intervene to address particularly inequitable outcomes? These are not easy questions to address. Local knowledge and contextual information critical for designing workable rules and policies are primarily centered among resource users. As a result, workable rules and policies require the active participation of resource users in governance. Empowering them, however, is fraught with challenges. Providing aid and assistance to local communities to bolster their conservation activities is not a straightforward process. Communities are not homogeneous political, social, and economic groupings that can be treated as a single unit (Agrawal and Gibson, 1999). External interventions, if not carefully crafted to the setting, may result in tragic unintended consequences.

1.02.3.2 Linking Water Uses and Administration across Multiple Scales and Jurisdictions How to characterize and think about the ties and linkages among governments, organizations, and groups, in order to gain traction for understanding and addressing practical problems is an ongoing struggle. The chasm between the dynamics of local-level, self-governing arrangements characterized by dense social networks and webs of norms, and central governments characterized by rent seeking, interest group politics, and corruption seems too vast to bridge. Conversely, the chasm between a relatively well-operating national government capable of exercising appropriate authority and deploying necessary resources and poorly designed local governments, or resource users who are unorganized and trapped in a race to harvest and who actively resist outside

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Governing Water: Institutions, Property Rights, and Sustainability

intervention appears to be insurmountable. However, the divide must be bridged to realize the promise of watershed governance. How do watershed and river basin governing systems get assembled, and assembled in ways that allow for the balancing of different values, accounting for externalities across users and jurisdictions, and the resolution of conflict in productive ways? Carlsson and Berkes (2005) propose a productive way of engaging with complex governing systems that span multiple scales: to focus initially on the functional aspects of governance, rather than the structural design of the governing institutions and organizations. More specifically, they propose a number of steps for analyzing problem-solving settings that draw heavily on the Institutional Analysis and Development Framework (IAD) developed by Ostrom and others (Kiser and Ostrom, 1982; Ostrom et al., 1994; Ostrom, 2007). First, what types of problems do water resources exhibit, or what problems and conflicts are water users experiencing? Second, once the social–ecological system has been identified and problems specified, including conflicting conceptualizations of problems, the policy analyst should turn to identifying who the participants are, how they are organized, and how they relate to and affect the essential management tasks implicated by the problems. As Carlsson and Berkes (2005: 73) explain, ‘‘The logic is that we start from the ‘bottom’, in the activities themselves, and try to figure out how management is organized, if power is shared, if rights and duties are contracted out, and if State authorities have ‘a finger in the pie’.’’ Third, how the actors and participants are linked, the types of organizational ties and types of authority they exercise that make them relevant to others in the situation, needs to be identified. Fourth, only then can remedies be sketched out that include strengthening governing capacity as well as more specific policy changes that address identified problems. The analytic steps that Carlsson and Berkes (2005) spell out overlap and interact; they do not constitute a strict linear process, they do not promise a particular set of outcomes, nor do they result in a specific configuration of institutional arrangements. Furthermore, if engaged in carefully, they require the policy analyst to grapple with a wide range of governance levels and issues, from the constitutional choice level, where the terms and conditions of governance are specified, to the collective-choice level where decision-making authority is exercised, and implementation and monitoring occur, to the operational level where actions and activities around water resource use take place (see also Ostrom, 2007). The value of using a problem-solving approach that incorporates different levels of governance is illustrated by comparing across the states of USA and the ability of water users, local jurisdictions, and state governments to assemble workable watershed-scale governing arrangements. At a first glance, it appears that among states whose constitutions grant local jurisdictions and water users considerable autonomy and decision-making authority, the authority is used to develop a variety of new associations, organizations, and governments at multiple scales that solve different problems. In addition, in investing in new organizations and governments, resource users expand their capacity to govern at the watershed scale. For instance, Landre and Travis (1998) describe the development of new forms of governance around the Keuka Lake and

watershed in the Finger Lakes region of New York. The lake’s water quality was slowly deteriorating, threatening a highly valued resource. The lake was used for drinking water and recreation (swimming, fishing, and boating). An association of homeowners surrounding the lake spearheaded the effort to protect the lake. Drawing on expertise from a local university and watershed planning grants from the state, the association launched a research and educational campaign, and developed a forum for public officials from the dozen or so surrounding towns. It was the public officials who had the authority to develop and implement water quality regulations. New York state permits local jurisdictions to enter into memoranda of agreement (MOA) by which local governments can jointly govern a shared resource or address a common problem. The local jurisdictions used the MOA to establish common septic tank regulation that applied to all homes, especially those surrounding the lake, and a septic tank inspector, to ensure that septic tanks were properly installed and maintained, thereby protecting the lake from pollution. The state could have intervened to address the water quality problem if local jurisdictions failed to act; however, local residents and governments chose to develop their own arrangements that they are responsible for and that are accountable to them (Keuka Lake Association, 2009). Much the same type of process has emerged among a number of watersheds in southern California, another state that grants citizens and local jurisdictions considerable authority and discretion to govern their water resources. As Schlager and Blomquist (2008) describe, the San Gabriel Basin covers much of Los Angeles County, one of the most highly urbanized counties in the country. Over 100 local political jurisdictions are located in the basin. The basin ties together these multiple jurisdictions through the San Gabriel River, the Rio Hondo River, three interconnected groundwater basins, and one groundwater basin not hydrologically connected to the others (Blomquist, 1992). Several water resource management problems have arisen in the San Gabriel River watershed, owing to the combined effects of the region’s limited water supplies, its extensive agricultural and then urban development, and the hydrogeology of the watershed itself. Each of these problems has been multi-jurisdictional in scope. Water users responded to each by developing new institutional arrangements. The arrangements are fitted together through a system of interorganizational and intergovernmental relationships. For instance, one of the initial problems addressed by local jurisdictions was importing water to meet the needs of rapidly growing populations and industries. A number of jurisdictions participated in the formation of the Metropolitan Water District in the 1920s to import water from the Colorado River Basin. As additional jurisdictions sought membership in the Metropolitan Water District and access to imported water supplies, they banded together and formed water districts whose purpose was to bring imported water to their member jurisdictions. Later, these districts participated in solving groundwater overdraft problems. Districts and larger municipalities spearheaded efforts to adjudicate rights in groundwater. To ensure representation of interests not adequately covered by municipalities and districts, water associations that encompassed the major water users within each basin were

Governing Water: Institutions, Property Rights, and Sustainability

formed and participated in developing agreements for allocating groundwater. Later still, the municipalities, districts, associations, and water masters, who monitor water rights, were the foundation on which water quality issues were addressed. Initially, water masters were given the task of developing water quality monitoring systems; however, as the discovery of water quality problems began to mount, and the projects and resources needed to remediate water supplies grew, a water quality authority was created to carry out the task of remediation. Thus, problem-by-problem water users and local jurisdictions assembled a San Gabriel River Basin governance system, sometimes granting new authorities to existing governments to address new problems, and sometimes creating new associations and governments to address problems. Watershed governance in both New York and California is predicated on allowing local jurisdictions and governments the authority to devise their own institutional solutions to shared problems. However, if water users and local governments did not want to participate for a multitude of reasons – they did not believe that (1) there was a problem, or (2) they were involved in causing the problem, or (3) they should pay to resolve the problem – they could not easily opt out or free ride off of the efforts of others. For instance, New York State could have intervened and established water quality standards for Keuka Lake and imposed regulations for meeting those standards. In southern California, local water user associations and districts, and municipal water providers regularly used state courts to bring all parties to the negotiating table to address water quantity problems. The California Health Department and the US Environmental Protection Agency, in setting water quality standards and in addressing highly polluted sites through the Superfund Program, called attention to and required action in relation to water quality problems. In other words, it is not just local autonomy that is important, but local autonomy and the larger institutional environment in which it is embedded that matters. Can local jurisdictions hold one another accountable, do they have access to conflict resolution mechanisms, and can they easily avoid participating in collective action to address water quantity or quality problems that they helped create? These are important considerations in watershed governance. The experience of other states suggests that it is no easy matter to go back and fill in institutional gaps in order to encourage, support, and require local jurisdictions to account for the impacts of their actions on others. For instance, until very recently, Nebraska water law and administration separated surface water from groundwater, even if the two are hydrologically linked (Schlager and Blomquist, 2008). Surface water, grounded in the prior appropriation doctrine, is administered by a state agency. Groundwater, grounded in the beneficial use doctrine, is administered by local natural resources districts, governed by groundwater users. The state had no legal authority to require natural resources districts to manage groundwater to account for externalities or environmental values. Beginning in the 1970s, with the listing of endangered species along the Platte River in south central Nebraska, planning and development of large surface water projects ceased (Aiken, 1999). By the 1990s, existing hydropower/irrigation projects were threatened, as the Federal

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Energy Regulatory Agency required the projects to account for their impacts on endangered species habitat to gain license renewal. Part of the problem the projects faced was the effects of groundwater pumping on their surface water supplies. In addition, Kansas, a downstream neighbor of Nebraska, filed a suit before the US Supreme Court, claiming that Nebraska was failing to abide by its legal obligations to allow Kansas’ share of the Republican River to flow through to Kansas because of groundwater pumping. Finally, Nebraska surface water rights holders began filing lawsuits against groundwater pumpers to try protect surface water flows. In each instance, the state of Nebraska was helpless to respond because it had no authority to require natural resources districts to regulate groundwater pumping. After two decades of conflict, that included multiple lawsuits, several governorappointed commissions, and numerous public hearings and meetings, a law was adopted that allowed the state water agency to declare river basins overappropriated. Such a designation would immediately trigger a well moratorium and require the natural resources districts within the overappropriated basin to develop groundwater regulations sufficient to bring the basin to a fully appropriated status. This is just the first, and most critical, step in developing ties and linkages among multiple governments and jurisdictions within Nebraska that will allow the coordination of ground and surface water and, which, in turn, will allow the state to meet its water obligations to surrounding states. Nebraska state law granted groundwater users significant authority to govern groundwater supplies through natural resource districts, but provided for few accountability mechanisms. The above cases highlight the value of taking a problemsolving approach to analyzing water resources issues and taking into account the ties and relations among citizens, organizations, and governments at multiple scales. As Young (2002: 266) argues, ‘‘The extent to which specific environmental or resource regimes yield outcomes that are sustainable – much less efficient or equitable – is a function not only of the allocation of tasks between or among institutions operating at different levels of social organization but also of crossscale interactions among distinct institutional arrangements.’’ Young (2002) provides a useful set of analytic concepts to assist policy analysts in assessing the institutional ties among water actors. The three analytic concepts are competence, compatibility, and capacity. Competence refers to the political and legal authority to engage in and implement commitments. For instance, in both New York and California, local jurisdictions possess the legal authority to engage in binding agreements with each other to regulate the use of a water source. However, the state of Nebraska did not have the competence to implement its agreement to provide Kansas with a designated amount of surface water. It did not possess such competence because it did not have the authority to regulate groundwater pumping. Compatibility refers to the congruence of institutional arrangements among governments operating in a given resource setting, such as a river basin. In the San Gabriel Basin, governing arrangements for each of the linked groundwater basins are now largely compatible and in line with California water doctrines; however, it took decades for those governing arrangements to be developed, with the downstream basin,

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Governing Water: Institutions, Property Rights, and Sustainability

which experienced problems first, taking the initial steps to regulate groundwater. As groundwater users in each of the upstream basins experienced problems and came to understand their own hydrologic ties, they too developed governing arrangements. The processes were adversarial and conflictual and took several decades to complete. In the end, each basin has its own groundwater rights and rules crafted to each setting, and the institutional arrangements are aligned so that complementarities and externalities are accounted for (Blomquist, 1992; Schlager and Blomquist, 2008). Capacity refers to the material resources and social capital needed to devise, implement, administer, and monitor rules and property rights systems. For instance, one of the challenges confronting the municipalities surrounding Keuka Lake was to provide sufficient resources to facilitate joint monitoring and enforcement of the septic tank rules adopted (Landre and Travis, 1998). A facilitator brokered an agreement that the costs would be shared equally among the jurisdictions and not be proportionate to the proximity of a jurisdiction to the lake. Each jurisdiction agreed to invest in the capacity of the MOA by making adequate resources available for its implementation. Competency, compatibility, and capacity further flesh out the problem-solving analytic approach that Carlsson and Berkes (2005) propose. One of the most important tasks in assessing a complex socio-ecological setting is identifying how the actors are linked, or not, and the quality of those linkages. Paying attention to the types and qualities of linkages also highlights the delicate balance that exists among different governments, associations, and organizations that constitute river basin or watershed governance. Just as it is challenging for regional and central governments to intervene and work with local communities of resource users without creating negative unintended consequences, so also is it difficult for actors to intervene in watershed governance systems to encourage, establish, and support productive linkages. However, what Carlsson and Berkes (2005), Young (2002), and many other scholars, who have attended to the governance of complex socio-ecological systems, have stressed is the importance of using a problem-centered approach for identifying capabilities and limitations of existing governance systems, and critically examining the competency, compatibility, and capacity of governments at different scales, rather than assuming that governments at specific levels are more competent or have greater capacity than governments at other levels (Larson, 2004; Ribot et al., 2006).

1.02.4 Courts: Hiding in Plain View As water users, water managers, and public officials attempt to link (or sever) ties among governments and organizations at different scales, they often contest the competency of a participant to engage in decision making; or they question the compatibility of different courses of action; or they seek ways to build the capacity of an organization, and in so doing they often turn to courts for assistance. Certainly, each of the US cases in the previous section include courts as an important actor in providing venues, shaping issues, and structuring solutions. Yet, in spite of talk about water governance revolving

around water conflicts and how to resolve them, the specific role of courts has been widely overlooked in both national and international contexts. In some countries, water experts recognize the practical importance of judicial decisions as features in the institutional landscape, but generally without thinking further about the courts’ significance. This is an unfortunate omission, because courts play a strategic and fundamental role in many countries’ political and economic systems. This is illustrated by a large academic literature, including comparative judicial politics (Jacob et al., 1996; Shapiro, 1981) and law and economics (e.g., Mercuro and Medema, 2006). The courts’ role is especially critical in situations of market-oriented policies and institutional frameworks. Market-driven governance aims to restrict state regulation, which requires strong judicial watchdogs. We return to this later in discussing the rule of law. Here, we intend to simply highlight a few key issues involving courts in the hope that people in water governance will be moved to investigate further. In the first place, courts and judges are one of the archetypal forms of conflict resolution. The judge sits apart and hears both sides of a dispute before deciding who wins and who loses. This is what political scientist Martin Shapiro has called the ‘‘logic of the triad in conflict resolution’’ (Shapiro, 1981: 1). Making that judicial decision means applying legal rules to specific fact situations, a process that requires reasoning, analysis, and interpretation. Particularly to people outside the legal profession, the judicial process is mysterious, and indeed how it works varies widely in different national and social contexts. In a classic book comparing different countries and legal traditions, Shapiro describes the conventional prototype of courts as growing out of conflict resolution. The prototype has four elements: ‘‘(1) an independent judge applying (2) preexisting legal norms after (3) adversary proceedings in order to achieve (4) a dichotomous decision in which one of the parties was assigned the legal right and the other found wrong’’ (Shapiro, 1981). Shapiro debunks each of these elements even in the context of conflict resolution, and he argues more broadly that courts also perform two other essential social functions: social control and lawmaking. The three functions often overlap. In the contexts of social control or lawmaking, however, the courts are farther away from their ‘‘basic social logic [and] perceived legitimacy.’’ Shapiro is worth quoting at length about the logic of courts: The basic social logic, or perceived legitimacy, of courts rests on the mutual consent of two persons in conflict to refer that conflict to a third for resolution. This basic logic is threatened by the substitution of office and law for mutual consent, both because one of the two parties may perceive the third as the ally of his enemy and because a third interest, that of the regime, is introducedy. When we move from courts as conflict resolvers to courts as social controllers, their social logic and their independence is even further undercut. For in this realm, while proceeding in the guise of triadic conflict resolver, courts clearly operate to impose outside interests on the parties. Finally, in the realm of judicial lawmaking, courts move furthest from their social logic and the conventional prototype because the rules they apply in the resolution of conflicts between two parties are neither directly consented to by the parties nor ‘preexisting.’ Instead, they are created by the third in the course of the conflict resolution itself. Thus, while the triadic mode of conflict resolution is nearly universal,

Governing Water: Institutions, Property Rights, and Sustainability

courts remain problematical in the sense that considerable tension invariably exists between their fundamental claims to legitimacy and their actual operations.’’ (Shapiro, 1981: 36–37, emphasis added)

The US is probably the most extreme example of a strong judicial role in water conflicts. The courts are major actors and decision makers in water policy and water rights and routinely combine the functions of conflict resolution, lawmaking, and social control. The courts are of course not the only important arena for addressing water conflicts, but they are a distinct arena with modes of reasoning and operation that are quite different from overtly political or economic approaches. Courts have been at the heart of one of the dominant slogans in international affairs in recent years: the rule of law. The idea of the rule of law shaped international debate about political and economic development in the 1980s and 1990s, as many countries went through historic processes of reform, generally toward democratization in the political sphere and toward markets in the economic sphere (Carothers, 1998; Dezalay and Garth, 2002; Thome, 2000; Trubek, 2006). This was approximately the same time that IWRM and sustainability came to dominate international debate about water. The rule of law means, in a nutshell, that laws apply to everyone, even powerful social actors and including government officials at all levels. Different views of the rule of law have put differing emphases on political versus economic issues. For many people, the rule of law refers to protection of human rights and democratic procedures of government. For others, it refers primarily to protection of economic and property rights from excessive government regulation. What all these views have in common is the goal of placing limits on the exercise of state power (Thompson, 1975; Whitlock, 2000). Moreover, all views agree that the ultimate watchdog is an independent judicial power, capable of challenging the legislative and executive powers of government. The courts are the foundation of the rule of law. The different political and economic aspects of the rule of law mean that courts are in a complicated position in many countries, and this bears directly on water conflicts. In political terms, the courts’ legitimacy is debatable because judges are not elected by the usual democratic processes and yet they sometimes overrule decisions made by popular sovereignty. In broader terms, social and economic as well as political, the courts are a critical arena for sorting through multiple values, rights, and interests. Many values are qualitatively different from each other and belong to different categories, and therefore cannot be quantitatively compared according to a single measure. Qualitative comparison is a task for legal and political institutions, and particularly courts in the case of conflicts. Law – whether made by legislators, executives, or judges – is vital to determining economic value, by creating the rules of the game and influencing how markets determine prices (Bromley, 1982; Commons, 1924; Whiteley et al., 2008).

1.02.5 Conclusion: Reconceptualizing Water Governance In an eye-catching headline ‘‘Stationarity is Dead: Whither Water Management?’’ a group of scientists argue that climate

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change is leading to changing hydrologic cycles and no longer can engineers rely on stationarity as the foundation for managing river basins (Milly et al., 2008). Stationarity refers to ‘‘the idea that natural systems fluctuate within an unchanging envelope of variability’’ (Milly et al., 2008: 573). Instead, the scientists argue, ‘‘nonstationary, probabilistic models of relevant environmental variables’’ must be developed to optimize water management. Furthermore, considerable attention must be paid to the rapid flow of information between climate scientists and water managers to make the new modeling efforts relevant to policy and management. How to realize such a complete transformation of the foundation of the design and management of large water projects? Interestingly, the authors propose a program that reflects the spirit of the Harvard Water Program. The Harvard Water Program, begun in the 1950s, used newly emerging computer technology and large data sets to illustrate the value and the possibility of computer simulations to examine alternative choices among objectives (Reuss, 1992, 2003). It provided a tool, multi-objective planning, for explicitly incorporating and analyzing multiple values and goals in developing and managing large water projects and the river basins in which they are situated. To that point in time, engineering considerations and values largely drove the design, planning, and operations process for developing water projects (Reuss, 1992). In proposing a new Harvard Water Program, Milly et al., to their credit, realize that responding to changing hydrologic cycles is not just an engineering problem, rather it is a societal problem. How river basins should be governed in light of climate change will entail engaging many people, organizations, associations, and governments expressing and pursuing many values. What, then, would a contemporary Harvard Water Program look like in what has been called the epoch of watershed sustainability (Sabatier et al., 2006)? The epoch of watershed sustainability is characterized by features both unique to it and in contrast to earlier time periods. It is characterized by many different stakeholder groups (interest groups, government agencies, scientists, local resource users, and native peoples) actively engaged in face-to-face interactions searching for win–win solutions to complex socio-ecological problems, grounded in extensive and intensive information development processes. In contrast, the Harvard Water Program existed during a time, in the US at least, when one federal agency had primary jurisdiction in a watershed or riverbasin, ‘‘with other agencies and interest groups acting as supplicants’’ (Sabatier et al., 2006: 23). The primary agency was ‘‘principally concerned with fulfilling its statutory mandate’’ with limited consideration given to other values (Sabatier et al., 2006: 24). The revolutionary aspect of the Harvard Water Program was a new methodology developed by experts for experts that allowed for the consideration of a wider range of values in river basin and water project planning and management. For a new Harvard Water Program to be successful, it would have to develop credible science, salient for a wide range of stakeholders, in processes viewed by scientists and stakeholders alike as legitimate (Cash et al., 2003). The challenge would no longer be how to fit a broader range of values into the design and operation of models used to optimize

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Governing Water: Institutions, Property Rights, and Sustainability

water development and management, as it was five decades ago, and as Milly et al. appear to envision it. Rather, the challenge would be how to design and fit organizations charged with the development of nonstationary, probabilistic hydrologic models that would be useful for water managers, policymakers, and citizens alike into complex, multiscale, intergovernmental, and organizational watershed governance systems. In other words, it would be a program that paid as much attention to institutional design (and all the disciplines that entails) as it would to the design of decision support systems (and all the disciplines that entails). IWRM and new modeling efforts around nonstationarity share much in common. Both recognize multiple values and the conflicts and politics that are likely to emerge as people fight, argue, contest, cooperate, and compete to ensure the realization of their cherished values. Rather than embracing such disorder, IWRM and new decision support models attempt to tidy things up by providing better information through models, or by bringing many interests and stakeholders to the table in search of common ground. While both approaches are useful, they fail to provide insight and guidance around water politics and governance as it unfolds in practice. What is instead needed and what we attempted to lay out in this chapter is an analytic approach for understanding complex social and ecological systems that recognizes a wide variety of institutional arrangements (including courts) and cross-scale linkages.

References Agrawal A and Gibson C (eds.) (1999) Communities and the Environment. New Brunswick, NJ: Rutgers University Press. Aiken JD (1999) Balancing endangered species protection and irrigation water rights: The Platte River cooperative agreement. Great Plains Natural Resources Journal 3: 119--158. Bauer C (2004) Siren Song: Chilean Water Law as a Model for International Reform. Washington, DC: RFF Press. Blomquist W (1992) Dividing the Waters: Governing Groundwater in Southern California. San Francisco, CA: ICS Press. Blomquist W, Schlager E, and Heikkila T (2004) Common Waters, Diverging Streams: Linking Institutions and Water Management in Arizona, California, and Colorado. Washington, DC: Resources for the Future. Bromley D (1982) Land and water problems: An institutional perspective. American Journal of Agricultural Economics 64(5): 834--844. Carlsson L and Berkes F (2005) Co-management: Concepts and methodological implications. Journal of Environmental Management 75: 65--76. Carothers T (1998) The rule of law revival. Foreign Affairs 77(2): 95--106. Cash D, Clark W, Alcock F, et al. (2003) Knowledge systems for sustainable development. Proceedings of the National Academy of Sciences of the United States of America 100: 8086--8091. Commons J (1924) Legal Foundations of Capitalism New York, NY: Macmillan. Conca K (2006) Governing Water: Contentious Transnational Politics and Global Institution Building. Cambridge, MA: MIT Press. Cosgrove W and Rijsberman F (2000) World Water Vision: Making Water Everybody’s Business. London: Earthscan. Coward EW (1979) Principles of social organization in an indigenous irrigation system. Human Organization 38: 28--36. Dezalay Y and Garth B (2002) Global Prescriptions: The Production, Exportation, and Importation of a New Legal Orthodoxy. Ann Arbor, MI: University of Michigan Press. Digital Library of the Commons (2009) Digital Library of the Commons Repository. http://dlc.dlib.indiana.edu/dlc (accessed April 2010). Dubash N, Dupar M, Kothari S, and Lissu T (2001) A Watershed in Governance? An Independent Assessment of the World Commission on Dams. World Resources Institute, Lokayan, and Lawyers Environmental Action Team.

GWP (Global Water Partnership) (2000a) Integrated Water Resources Management, TAC Background Paper No. 4. Stockholm: Global Water Partnership. GWP (Global Water Partnership) (2000b) Towards Water Security: A Framework for Action. Stockholm: Global Water Partnership. Hardin G (1968) The tragedy of the commons. Science 162: 1243--1248. Ilahiane H (1999) The ethnopolitics of irrigation management in the Ziz Oasis, Morocco. In: Agrawal A and Gibson C (eds.) Communities and the Environment: Ethnicity, Gender, and the State in Community-Based Conservation, pp. 89--110. New Brunswick, NJ: Rutgers University Press. Jacob H, Blankenburg E, Kritzer HM, Provine DM, and Sanders J (1996) Courts, Law, and Politics in Comparative Perspective. New Haven, CT: Yale University Press. Kemper KE, Blomquist W, and Dinar A (eds.) (2007) Integrated River Basin Management through Decentralization. Berlin: Springer. Keuka Lake Association (2009) About Keuka Lake Association. http:// www.keukalakeassoc.org/what/about_kla.php (accessed April 2010). Kiser L and Ostrom E (1982) The three worlds of action: A metatheoretical synthesis of institutional approaches. In: Ostrom E (ed.) Strategies of Political Inquiry, pp. 179--222. Beverly Hills, CA: Sage. Lam WF (1998) Governing Irrigation Systems in Nepal: Institutions, Infrastructure, and Collective Action. San Francisco, CA: ICS Press. Landre P and Travis L (1998) Collaborative watershed management in the Finger Lakes region, New York. Presented at the Conference of the International Association for the Study of Common Property. Vancouver, BC, 10–14 June 1998. http:// www.indiana.edu/˜iascp/Final/landre.pdf (accessed April 2010). Lansing JS (1991) Priests and Programmers: Technologies of Power in the Engineered Landscape of Bali. Princeton, NJ: Princeton University Press. Larson A (2004) Formal decentralization and the imperative of decentralization ‘from below’: A case study of natural resource management in Nicaragua. European Journal of Development Research 16(1): 55--70. Maass A and Anderson RL (1986) y and the Desert Shall Rejoice: Conflict, Growth, and Justice in Arid Environments. Malabar, FL: R.E. Krieger. Mercuro N and Medema S (2006) Economics and the Law, 2nd edn. Princeton, NJ: Princeton University Press. Milly PC, Betancourt J, Falkenmark M, et al. (2008) Stationarity is dead: Whither water management? Science 319: 573--574. Olson M (1965) The Logic of Collective Action: Public Goods and the Theory of Groups. Cambridge, MA: Harvard University Press. Ostrom E (1990) Governing the Commons: The Evolution of Institutions for Collective Action. New York, NY: Cambridge University Press. Ostrom E (1999) Coping with the tragedy of the commons. Annual Review of Political Science 2: 493--535. Ostrom E (2007) Understanding Institutional Diversity. Princeton, NJ: Princeton University Press. Ostrom E, Dietz T, Dolsak N, et al. (2002) The Drama of the Commons. Washington, DC: National Academies Press. Ostrom E, Gardner R, and Walker J (1994) Rules, Games, and Common Pool Resources. Ann Arbor, MI: University of Michigan Press. Reuss M (1992) Coping with uncertainty: Social scientists, engineers, and federal water resources planning. Natural Resources Journal 32: 101--135. Reuss M (2003) Is it time to resurrect the Harvard Water Program? Journal of Water Resources Planning and Management 129(5): 357--360. Ribot J, Agrawal A, and Larson A (2006) Recentralizing while decentralizing: How national governments reappropriate forest resources. World Development 34(11): 1864--1886. Rogers P (2002) Water Governance in Latin America and the Caribbean. Washington, DC: Inter-American Development Bank. Rogers P and Hall A (2003) Effective Water Governance, TAC Background Paper No. 7. Stockholm: Global Water Partnership. Rose C (2002) Common property, regulatory property, and environmental protection: Comparing community-based management to tradable environmental allowances. In: Ostrom E, Dietz T, Dolsak N, et al. (eds.) Drama of the Commons, pp. 263--292. Washington, DC: National Academies Press. Sabatier P, Weible C, and Ficker J (2006) Eras of water management in the United States: Implications for collaborative watershed approaches. In: Sabatier P, Focht W, Lubbell M, et al. (eds.) Swimming Upstream: Collaborative Approaches to Watershed Management, pp. 23--52. Cambridge, MA: MIT Press. Schlager E (2005) Getting the relationships right in water property rights. In: Bruns BR, Ringler C, and Meinzen-Dick R (eds.) Water Rights Reform: Lessons for Institutional Design, pp. 27--54. Washington, DC: International Food Policy Research Institute. Schlager E and Blomquist W (2005) Beneath the surface: Shared attributes of fisheries and aquifers and implications for institutional design. Prepared for Presentation at

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Festschrift for Elinor Ostrom, Workshop on Political Theory and Policy Analysis, Indiana University, Bloomington, IN, 22–23 November 2005. Schlager E and Blomquist W (2008) Embracing Watershed Politics. Boulder, CO: University Press of Colorado. Schlager E and Ostrom E (1992) Common property and natural resources: A conceptual analysis. Land Economics 68: 249--252. Shapiro M (1981) Courts: A Comparative and Political Analysis. Chicago, IL: University of Chicago Press. Tang SY (1992) Institutions and Collective Action: Self-Governance in Irrigation. San Francisco, CA: ICS Press. Tang SY (1994) Institutions and performance in irrigation systems. In: Ostrom E, Gardner R, and Walker J (eds.) Rules, Games, and Common Pool Resources, pp. 225--246. Ann Arbor, MI: University of Michigan Press. Thome J (2000) Heading south but looking north: Globalization and law reform in Latin America. Wisconsin Law Review 2000: 691--712. Thompson EP (1975) Whigs and Hunters: The Origin of the Black Act. London: Allen Lane.

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Trubek D (2006) The ‘rule of law’ in development assistance: Past, present, and future. In: Trubek DM and Santos A (eds.) The New Law and Economic Development: A Critical Appraisal, pp. 74--94. New York, NY: Cambridge University Press. Whiteley J, Ingram H, and Perry R (2008) Water, Place and Equity. Cambridge, MA: MIT Press. Whitlock W (2000) The rule of law. Wisconsin Law Review 2000: 723--742. World Bank (1993) Water Resources Management: A World Bank Policy Paper. Washington, DC: World Bank. World Bank (2004) Water Resources Sector Strategy: Strategic Directions for World Bank Engagement. Washington, DC: World Bank. World Commission on Dams (2000) Dams and development: A new framework for decision-making. The Report of the WCD. London: Earthscan. Young O (2002) Institutional interplay: The environmental consequences of cross-scale interactions. In: Ostrom E, Dietz T, Dolsak N, et al. (eds.) Drama of the Commons, pp. 263--292. Washington, DC: National Academies Press.

1.03 Managing Aquatic Ecosystems CM Finlayson, Charles Sturt University, Albury, NSW, Australia & 2011 Elsevier B.V. All rights reserved.

1.03.1 1.03.2 1.03.2.1 1.03.2.2 1.03.2.3 1.03.3 1.03.3.1 1.03.3.2 1.03.3.3 1.03.3.4 1.03.4 1.03.4.1 1.03.4.2 1.03.4.3 1.03.4.4 1.03.4.5 1.03.4.6 1.03.5 1.03.5.1 1.03.5.2 1.03.5.3 1.03.5.4 1.03.6 References

Introduction Key Concepts Wise Use of Wetlands Ecological Character The Ecosystem Approach Distribution and Classification of Aquatic Ecosystems Classification of Inland Aquatic Ecosystems Extent and Distribution Loss and Degradation of Inland Aquatic Ecosystems Loss of Species from Inland Aquatic Ecosystems Drivers of Change in Inland Aquatic Ecosystems Drainage, Clearing, and Infilling Modification of Water Regimes Invasive Species Overfishing Water Pollution and Eutrophication Climate Change Management Responses Integrated Management Processes International Cooperation and Action Restoration and Wise Use of Wetlands Supporting Local Community Involvement in Management Conclusions

1.03.1 Introduction The importance of aquatic ecosystems for people has been highlighted in recent years by the Millennium Ecosystem Assessment (Finlayson et al., 2005) and subsequent assessments such as the World Water Development Report (UNESCOWWAP, 2006), the Global International Waters Assessment (UNEP, 2006), the Global Biodiversity Outlook (Secretariat of the Convention on Biological Diversity, 2006) and the Global Environment Outlook (UNEP, 2007), and the Comprehensive Assessment of Water Management in Agriculture (Molden, 2007). However, despite many aquatic ecosystems being highly important for people and biodiversity, they have been degraded over many decades and many lost (Finlayson and D’Cruz, 2005). The continued degradation and loss of these highly valued ecosystems bring into question the effectiveness of current management practices and whether or not different approaches should be developed to ensure that they are managed wisely. The scope and definition of aquatic ecosystems have been widely discussed with many different definitions and classifications being used to cover a range of inland and coastal aquatic ecosystems or wetlands (Finlayson and van der Valk, 1995; Mitsch and Gosselink, 2007; Whigham, 2009). The term wetland has often been used to define a narrow range of inland aquatic systems, such as bogs, marshes, and swamps, while at other time it has been used to define a wider range

35 35 35 37 39 42 42 43 44 45 46 47 48 50 52 53 54 55 55 56 56 56 56 57

including rivers, lakes, reservoirs, and rice fields as well (Finlayson et al., 1999). Extensive information on wetland definition and delineation is available (e.g., Finlayson and van der Valk, 1995; Mitsch et al., 1994), but the failure to consider the different definitions completely that have been used around the world has resulted in confusion and inaccurate analyses on the extent and condition of these ecosystems (Finlayson and Spiers, 1999). The term aquatic ecosystem is used here in order to illustrate that the wider range of inland wetlands is being considered. This corresponds with the approach taken in the Millennium Ecosystem Assessment and encompasses the terms inland water systems, inland waters, or inland wetlands and includes marshes, swamps, lakes, and rivers, regardless of their size or whether they are permanent or temporary or saline or fresh (Finlayson and D’Cruz, 2005). As there is no clear boundary between inland and coastal aquatic ecosystems, the differentiation between them is indicative only. As there are strong interactions between inland and coastal aquatic ecosystems, a clear differentiation is not adopted. Coastal brackish and saline marshes, estuaries, and mangroves are not considered unless there is a strong connection between them and nearby freshwater aquatic ecosystems. Inland salt lakes are included, especially as many undergo flooding, whether periodic or irregular, with freshwater. The range of ecosystems considered is shown in Figure 1.

35

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Figure 1 Inland aquatic ecosystems.

The text below outlines some of the key concepts and approaches for managing inland aquatic ecosystems with a focus on water management. Key sources of information include material derived from the Ramsar Convention on Wetlands (Ramsar Convention Secretariat, 2006a), the Millennium Ecosystem Assessment (Finlayson et al., 2005), and the Wetland Handbook (Maltby and Barker, 2009). The importance of the Ramsar Convention in promoting the conservation and wise use of wetlands globally cannot be underestimated – the text of the Convention was signed in 1971 and in many ways was a forerunner of many subsequent developments in conservation, especially the shift from species preservation to sustainable use and integrated management of wetland ecosystems (Matthews, 1993). A summary of the key features of the Convention is provided in Box 1.

1.03.2 Key Concepts A number of key concepts associated with the management of inland aquatic ecosystems are outlined in the section below. These include the concepts of wise use of wetlands and ecological character of wetlands used by the Ramsar Convention on Wetlands, and the ecosystem approach used by the

Convention on Biological Diversity and promoted by the IUCN Commission on Ecosystem Management.

1.03.2.1 Wise Use of Wetlands Since its inception in 1971, the Convention on Wetlands (Ramsar, Iran, 1971) has promoted the wise use of wetlands (Matthews, 1993). The contracting parties to the Convention have subsequently accepted a mission statement that commits them to ‘‘y the conservation and wise use of all wetlands through local, regional and national actions and international cooperation, as a contribution towards achieving sustainable development throughout the world’’ (Ramsar Strategic Plan 2009–2015). The contracting parties have also agreed to deliver this mission through three streams of activities: (1) the wise use of all wetlands, (2) the designation and management of wetlands of international importance (Ramsar sites; see Figure 2), and (3) international cooperation. Wise use of wetlands was included in the text of the Convention under article 3.1 ‘‘The Contracting Parties shall formulate and implement their planning so as to promote y as far as possible the wise use of wetlands in their territory.’’ Matthews (1993) reported that from the outset the wise use of wetlands was seen as the maintenance of their ecological character, as a basis not only for nature conservation, but for

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37

Box 1 The Ramsar Convention on Wetlands. Information from The Ramsar Convention on Wetlands, http://www.ramsar.org (accessed September 2010). The Convention on Wetlands of International Importance, called the Ramsar Convention, is an intergovernmental treaty that provides a framework for national action and international cooperation for the conservation and wise use of wetlands and their resources. Unlike the other global environmental conventions, Ramsar is not affiliated with the United Nations system of Multilateral Environmental Agreements, but it works very closely with the other biodiversity-related treaties and agreements. The Convention was negotiated through the 1960s by countries and nongovernmental organizations that were concerned at the increasing loss and degradation of wetland habitat for migratory waterbirds; the treaty was adopted in the Iranian city of Ramsar in 1971 and came into force in 1975. It is the only global environmental treaty that deals with a particular ecosystem. The Convention’s mission is ‘‘the conservation and wise use of all wetlands through local and national actions and international cooperation, as a contribution towards achieving sustainable development throughout the world.’’ It uses a broad definition of the types of wetlands covered in its mission, including lakes and rivers, swamps and marshes, wet grasslands and peatlands, oases, estuaries, deltas and tidal flats, near-shore marine areas, mangroves and coral reefs, and human-made sites such as fish ponds, rice paddies, reservoirs, and salt pans. At the center of the Convention is the wise use concept which has at its heart the conservation and sustainable use of wetlands and their resources, for the benefit of human kind. The Convention has 160 contracting parties (12 December 2009) that have agreed to four main obligations: 1. designate at least one wetland at the time of accession for inclusion in the list of wetlands of international importance (the Ramsar List) and to promote its conservation, and in addition to continue to designate suitable wetlands within its territory for the list; 2. include wetland conservation in their national land-use planning and to promote, as far as possible, the wise use of wetlands in their territory; 3. establish nature reserves in wetlands, whether or not they are included in the Ramsar List, and to promote training in the fields of wetland research, management, and wardening; and 4. consult with other contracting parties about implementation of the Convention, especially in regard to transboundary wetlands, shared water systems, and shared species.

There are currently 1891 wetland sites (information as of 23 July 2010) designated for the list of wetlands of international importance covering a total surface area of 185 464 092 ha (Figure 2). Further information on the history and development of the Convention can be obtained from Matthews (1993).

Figure 2 Wetlands of international importance listed under the Ramsar Convention on Wetlands. From http://www.ramsar.org.

sustainable development also. With this background, wise use of wetlands was later defined as ‘‘y their sustainable utilisation for the benefit of humankind in a way compatible with the maintenance of the natural properties of the ecosystem.’’ With sustainable utilization it is defined as ‘‘y human use of a wetland so that it may yield the greatest continuous benefit to present generations while maintaining its potential to meet the needs and aspirations of future generations’’ (Davis, 1993).

Guidelines for the wise use of wetlands were formally adopted by the Convention in 1990 and emphasized the link between wetlands and people, and placed wetlands in a catchment or coastal zone context (Davis, 1993; Ramsar Convention Secretariat, 2006b; Maltby, 2009). The definition and guidelines remained in place until 2005 when wise use was redefined as ‘‘y the maintenance of their ecological character, achieved through the implementation of ecosystem approaches, within the context of sustainable development.’’

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The change in definition strengthened the link between wise use and the concept of sustainable development that had been promoted by the UN World Commission on Environment and Development (Brundtland, 1987). The change in definition also reflected an expansion in the wise use guidelines with the development and adoption of an increasing range of policy and technical guidelines known as the Ramsar Toolkit of Wise Use Handbooks (Ramsar Convention Secretariat, 2006c). Table 1 provides a summary of the material covered in the handbooks. The guidelines have also been mapped onto the conceptual framework developed for the Millennium Ecosystem Assessment which is consistent with the Ramsar concept of wise use (Finlayson et al., 2005). The framework provides a guide to the Ramsar wise use guidelines, and can be used to identify gaps (Figure 3). Many of the guidelines apply directly to ecosystem

services and the links between these services and the ecological components and processes that characterize wetlands, including those for describing and assessing the condition of the wetland. Others address interventions covering the direct drivers of change to ecosystems while two sets – those covering national wetland policies and on reviewing legislative and institutional frameworks – deal with the indirect drivers of change. Some, such as those on international cooperation, communications, education, and public awareness, apply to several parts of the framework. The guidelines are under continual review by the Convention and updated at regular intervals. It is anticipated that further attention will be devoted to guidelines that address policy issues and encourage greater integration between resource users and conservation interests – themes that have been evident in recent global assessments covering biodiversity and ecosystem services.

Table 1 Guidance available through the 3rd edition of the Ramsar Wise Use Handbooks and the relevant Resolutions agreed by the Contracting Parties to the Convention Handbook no.

Title

Content

Resolutions

1.

Wise use of wetlands

IX.1

2. 3.

National wetland policies Laws and institutions

A Conceptual Framework for the wise use of wetlands and the maintenance of their ecological character Developing and implementing National Wetland Policies

4.

Wetland CEPA

5.

Participatory skills

6.

Water-related guidance River basin management Water allocation and management Managing groundwater Coastal management Inventory, assessment, and monitoring Water allocation and management Impact assessment

7. 8. 9. 10. 11.

12.

Reviewing laws and institutions to promote the conservation and wise use of wetlands The Convention’s Programme on communication, education, and public awareness (CEPA) 2003–2008 Establishing and strengthening local communities’ and indigenous people’s participation in the management of wetlands An integrated framework for the Convention’s water-related guidance Integrating wetland conservation and wise use into river basin management Guidelines for the allocation and management of water for maintaining the ecological functions of wetlands Guidelines for the management of groundwater to maintain wetland ecological character Wetland issues in Integrated Coastal Zone Management An integrated framework for wetland inventory, assessment, and monitoring

14.

Designating Ramsar sites

15.

Addressing change in ecological character Managing wetlands

Guidelines for the allocation and management of water for maintaining the ecological functions of wetlands Guidelines for incorporating biodiversity-related issues into environmental impact assessment legislation and/or processes and in strategic environmental assessment The Strategic Framework and guidelines for the future development of the List of Wetlands of International Importance Addressing change in the ecological character of Ramsar sites and other wetlands Frameworks for managing Ramsar sites and other wetlands

International cooperation

Guidelines for international cooperation under the Ramsar Convention on Wetlands

13.

16. 17.

VII.6 VII.7 VIII.31 VII.8

IX.1 VII.19, IX.1, IX.3 VIII.1 IX.1 VII.21, VIII.4 IX.1

VIII.1 VII.16, VIII.9

VIII.10

V.4, VI.1, VII.24, VIII.8, VIII.16, VIII.20, VIII.22, IX.6 V.7, VI.1, VII.10, VIII.14, VIII.18, VIII.19, IX.4 VII.19

Reproduced from The Ramsar ‘Toolkit’, 3rd edn. (2007) The Ramsar Handbooks for the Wise Use of Wetlands. http://www.ramsar.org/cda/ramsar/display/main/ main.jsp?zn=ramsar&cp=1-30^21323_4000_0 (accessed August 2010).

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39

Global Regional Local Human well-being and poverty reduction

Indirect drivers of change

• Health security • Environmental security • Economical security • Cultural security • Equity

• Demographic • Economic (e.g., trade, subsides, markets) • Sociopolitical(e.g., governance, institutional and legal framework) • Science and technology • Cultural and religious (e.g., choices about what and how much to consume)

777

171 HB17: International cooperation

HB0: Water allocation and management

HB1: Wise use

HB3:Laws and institutions

HB4: Wetland CEPA HB7: River basin Management

171

HB16: Managing wetlands HB12: Wetland inventory

Ecosystem services

HB11: Inventory, assessment, monitoring

HB14: Designating rumuer sites HB5: Participatory skills

HB10: Central management

• Provisioning (e.g., food, fresh water, fuel, genetic resources) • Regulating (e.g., climate, water, natural hazard mitigation) • Cultural (e.g., spiritual, aesthetic) • Supporting (e.g., primary production, nutrient cycling)

Life on earth: biodiversity Strategies and Interventions

HB2: National wetlands policies

Direct drivers of change

HB13: Impact assessment

HB9: Groundwater

HB8: Water allocation and management

• Changes in local land use and land cover • Species removals and/or invasive introductions • Eutrophication and pollution • Hydraulic infrastructure development • Water abstraction • Climate change HBO: Water-related guidance

HB10: Coastal management

HB7: River basin management HB15: Change in cool character

777 No specific guidance

HBxx

Dark background: Handbooks include interventions into several red bars

Figure 3 A framework for the wise use of wetlands and the application of the guidelines in the Ramsar ‘Toolkit’ of Wise Use Handbooks, 3rd edn. Reproduced from Ramsar Convention Secretariat (2006b) Wise use of wetlands: A conceptual framework for the wise use of wetlands. In: Ramsar Handbooks for the Wise Use of Wetlands, 3rd edn., vol. 1. Gland, Switzerland: Ramsar Convention Secretariat, with permission from Ramsar; updated from Finlayson CM, D’Cruz R, and Davidson NJ (2005) Ecosystem Services and Human Well-Being: Water and Wetlands Synthesis. Washington, DC: World Resources Institute.

1.03.2.2 Ecological Character The concept of ecological character was introduced in the text of the Ramsar Convention and is now seen as a basis for the wise use of wetlands globally. Contracting parties to the Ramsar Convention are required to promote the conservation of all wetlands through the maintenance of their ecological character and to do this they are expected to establish management planning and monitoring mechanisms (Ramsar Convention Secretariat, 2006a). Ecological character is now defined as ‘‘y the combination of the ecosystem components, processes and benefits/services that characterise the wetland at a given point in time’’ (Ramsar Convention Secretariat, 2007). The previous definition considered ecosystem services separately to the ecological components and processes that were seen as comprising the ecological character of a wetland (Figure 4). Ecosystem services were incorporated into the definition of ecological character after the Convention adopted the findings of the Millennium Ecosystem Assessment as they applied to water and wetlands (Finlayson et al., 2005). This widened the concept of wetland conservation and provided an overt link with the management and uses of a wetland, as expressed through ecosystem services which provide benefits to people.

The description of the ecological character of a wetland provides baseline data that establish the range of natural variation in ecological components and processes and ecosystem services at each site within a given time frame, against which change can be assessed. To describe the ecological character of a wetland and ascertain when adverse change may have occurred, the Convention has established an integrated framework for wetland inventory, assessment, and monitoring (Ramsar Convention Secretariat, 2006c; Finlayson et al., 2005). In support of this approach, the Convention also adopted the concept of ecosystem services as provided by the Millennium Ecosystem Assessment (2003) with four categories of services (Figure 5): 1. provision services – products obtained from ecosystems; 2. regulating services – benefits obtained from regulation of ecosystem processes; 3. cultural services – nonmaterial benefits obtained from ecosystems; and 4. supporting services – services necessary for the production of all other ecosystem services.

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Provisioning Biological Regulating

Chemical Physical

Cultural

Supporting

Ecosystem components and processes

Ecosystem services

Figure 4 The ecological character of a wetland showing the relationship between the ecosystem components and processes and services that comprise the wetland.

Provisioning services

Regulating services

Cultural services

Goods produced or provided by ecosystems

Benefits from regulation of ecosystem processes

Nonmaterial benefits from ecosystems

• Food • Fuel wood • Fiber • Timber

• Water partitioning • Pest regulation • Climate regulation • Pollination

• Spiritual • Recreational • Aesthetic • Educational

Support services Factors necessary for producing ecosystem services • Hydrological cycle • Soil formation • Nutrient cycling • Primary production

Figure 5 Four categories of ecosystems services as outlined in the Millennium Ecosystem Assessment (2003). Reproduced from Millennium Ecosystem Assessment (2003) Ecosystems and Human Well-Being: A Framework for Assessment. Washington, DC: Island Press.

The relative importance of these services in different types of inland aquatic ecosystems is shown in Figure 6. The information in this figure was based on expert opinion given the absence of sufficient data to support a quantitative analysis. Given this situation the collection of further information on the extent of ecosystem services provided by wetlands is encouraged by the Convention as a basis for wise use.

1.03.2.3 The Ecosystem Approach In response to widely articulated weaknesses in sectoral approaches for wetland management (e.g., Hollis, 1992, 1998), a number of ecosystem approaches have been developed to promote the conservation and sustainable and equitable use of wetlands (Brown et al., 2005). The Convention on

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41

Alpine and tundra wetlands

Springs and oases

Geothermal wetlands

Underground wetlands including caves and groundwater systems

Seasonal lakes, marshes, and swamps including floodplains Forested wetlands, marshes, and swamps, including floodplains

Permanent lakes, reservoirs

Permanent and temporary rivers and streams

Comments and examples

Services

Scale is low •, medium , to high ; not known?; blank cells indicate that the service is not considered applicable to the wetland type. The information in the table represents expert opinion for a global average pattern for wetlands; there will be local and regional differences in relative magnitudes.

?

?

?

?

?

?

?

?

?

?

Inland wetlands Provisioning Food Fresh water

Fiber and fuel Biochemical products Genetic materials

Regulating Climate regulation

Hydrological regimes

Pollution control and detoxification Erosion protection

Natural hazards Cultural Spiritual and inspirational Recreational Aesthetic Educational Supporting Biodiversity Soil formation Nutrient cycling Pollination

Production of fish, wild game, fruits, grains, and so on Storage and retention of water; provision of water for irrigation and for drinking Production of timber, fuelwood, peat, fodder, aggregates Extraction of materials from biota Medicine; genes for resistance to plant pathogens, ornamental species, and so on

? ?

?

Regulation of greenhouse gases, temperature, precipitation, and other climatic processes; chemical composition of the atmosphere Groundwater recharge and discharge; storage of water for agriculture or industry Retention recovery, and removal of excess nutrients and pollutants Retention of soils and prevention of structural change (such as coastal erosion, bank slumping, and so on) Flood control; storm protection

?

Personal feelings and well-being; religious significance Opportunities for tourism and recreational activities Appreciation of natural features Opportunities for formal and informal education and training Habitats for resident or transient species Sediment retention and accumulation of organic matter Storage, recycling, processing, and acquisition of nutrients

?

Support for pollinators

Figure 6 Ecosystem services provided by different inland aquatic ecosystems. Reproduced from Finlayson CM, D’ Cruz R, and Davidson NJ (2005) Ecosystem Services and Human Well-Being: Water and Wetlands Synthesis. Washington, DC: World Resources Institute; and Millennium Ecosystem Assessment (2005) Ecosystems and Human Well-Being: Synthesis. Washington, DC: Island Press.

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Table 2 1. 2. 3. 4.

5. 6. 7. 8. 9. 10. 11. 12.

Principles of the ecosystem approach adopted by the Convention on Biological Diversity

The objectives of management of land, water, and living resources are a matter of societal choices Management should be decentralized to the lowest appropriate level Ecosystem managers should consider the effects (actual or potential) of their activities on adjacent and other ecosystems After recognizing potential gains from management, there is a need to understand the ecosystem in an economic context. Any ecosystem management program should: (a) reduce those market distortions that adversely affect biological diversity; (b) align incentives to promote sustainable use; and (c) internalize costs and benefits in the given ecosystem to the extent feasible Conservation of ecosystem structure and functioning, in order to maintain ecosystem services, should be a priority target of the ecosystem approach Ecosystems must be managed within the limits to their functioning The ecosystem approach should be undertaken at the appropriate spatial and temporal scales Recognizing the varying temporal scales and lag effects which characterize ecosystem processes, objectives for ecosystem management should be set for the long term Management must recognize that change is inevitable The ecosystem approach should seek the appropriate balance between, and integration of, conservation and use of biological diversity The ecosystem approach should consider all forms of relevant information, including scientific, indigenous, and local knowledge, innovations, and practices The ecosystem approach should involve all relevant sectors of society and scientific disciplines

Information derived from Ecosystem Approach, Principles, http://www.cbd.int/ecosystem/principles.shtml (accessed August 2010).

Biological Diversity has promoted the ecosystem approach by focusing on managing environmental resources and by promoting a balance between human needs and biodiversity. The wise use approach adopted by the Ramsar Convention as well as integrated coastal zone management and integrated catchment management are compatible with the ecosystem approach adopted by the Convention on Biological Diversity (Shepherd, 2004; Brown et al., 2005). The ecosystem approach adopted by the Convention on Biological Diversity was seen as a way of reaching a balance between the objectives of the Convention covering conservation, sustainable use, and the fair and equitable sharing of the benefits arising out of the utilization of genetic resource. It is based on a strategy for the integrated management of land, water, and living resources and promotes conservation and sustainable use in an equitable way (Shepherd, 2004). It places human needs at the center of biodiversity management and acknowledges that ecosystems perform multiple functions of importance to people both locally and further afield. The guiding principles and strategies adopted by the Convention are shown in Table 2. As the Convention’s ecosystem approach does not comprise a specific applicable method, it has been criticized for being too vague to be of practical value, while others have highlighted its flexibility (Brown et al., 2005). Taking note of the criticisms, Shepherd (2004) grouped the guiding principles into the following five steps, each involving a range of actions to encourage discussion, planning, and step-by-step action:

1. determine the main stakeholders, define the ecosystem area, and develop the relationship between them; 2. characterize the structure and function of the ecosystem, and set in place mechanisms to manage and monitor it; 3. identify the important economic issues that will affect the ecosystem and its inhabitants; 4. determine the likely impact of the ecosystem on adjacent ecosystems; and 5. decide long-term goals and flexible ways of reaching them.

Brown et al. (2005) have highlighted some of the constraints that are generally seen to characterize ecosystem approaches. These include: a failure to consider specific areas, resources, or species that may need a more targeted approach for their conservation; uncertainties and lack of guidance about how to balance conservation and sustainable use; and difficulties with establishing collaboration between stakeholders and negotiating trade-offs between them in a fair and equitable way. Even with these constraints the principles outlined in Table 2 have been widely accepted and, in one way or the other, now feature in many current approaches for managing aquatic ecosystems (Finlayson and D’Cruz, 2005; Maltby, 2009). The principles adopted by the Convention on Biological Diversity are generally applicable to other ecosystem approaches and have largely been addressed through the more detailed guidance for wise use developed by the Ramsar Convention (Ramsar Convention Secretariat, 2006a, 2006b). Ecosystem approaches generally provide for a broad, crosssectoral approach for managing aquatic ecosystems by addressing both direct and indirect drivers of change and considering the multiple benefits people derive from the ecosystem services that they provide.

1.03.3 Distribution and Classification of Aquatic Ecosystems The extent and distribution of inland aquatic ecosystems is poorly and unevenly known at the global and regional scales due to differences in definitions as well as difficulties in delineating and mapping ecosystems with variable boundaries due to fluctuations in water levels (Finlayson et al., 1999; Rebelo et al., 2009; Mackay et al., 2009). In many cases, comprehensive documentation of the extent and distribution of inland aquatic ecosystems at the regional or national levels also does not exist. The larger ecosystems, such as lakes and inland seas, have been mapped along with the major rivers, but for many parts of the world smaller ecosystems are not well mapped or delineated. As a consequence, assessment of

Managing Aquatic Ecosystems

the extent of and change in these ecosystems at the continental level is compromised by the inconsistency and unreliability of the data.

1.03.3.1 Classification of Inland Aquatic Ecosystems The classification of inland aquatic ecosystems or wetlands has consumed an inordinate amount of time and controversy with many systems developed and used in different countries (see summary in Finlayson and van der Valk (1995)). While a wetland inventory can be undertaken without recourse to an agreed classification, given that a standardized and logical process of data collection or collation is undertaken, it is likely to become necessary to classify them at some stage during the assessment phase of wetland management, especially where it is useful or necessary to make comparisons between different wetlands. At this stage an agreed set of terms is not only desirable but possibly mandatory to ensure conformity of comparisons and hence decisions. Thus, the importance of classification cannot be overstated, but it equally needs to be Table 3

remembered that classification is a tool within a larger set of tools that are designed to provide an adequate information base for the wise use, conservation, and management of all wetlands. The Ramsar wetland definition is supported by a classification scheme with 42 categories that is purposefully simple and global in scope, and readily compatible with other classifications that may be preferred regionally, nationally, or locally (Scott and Jones, 1995). The classification divides wetlands into three main types: marine/coastal (12 wetland types), inland (20), and human-made (10), with several categories, according to vegetation, soil/rock, inundation, water quality (freshwater to saline water), and landform, within each type (Ramsar Convention Secretariat, 2006d). The characteristics of the inland wetland types are shown in Table 3. While used in a general sense by contracting parties to the Convention, it has also been augmented by more specific classifications. The broad wetland categories included in the classification have been particularly useful when comparing between countries or regions, but less so when more specific

Characteristics of inland wetland types contained within the Ramsar wetland classification

Wetland type

Wetland characteristics

1.

Permanent rivers/streams/creeks

2. 3. 4.

Permanent inland deltas Freshwater springs; oases Seasonal/intermittent/irregular rivers/ streams/creeks

5. 6. 7. 8.

Permanent freshwater lakes Permanent freshwater marshes/pools Seasonal/intermittent freshwater lakes Seasonal/intermittent freshwater marshes/ pools on inorganic soils

Lakes and pools

9. 10. 11. 12.

Permanent freshwater marshes/pools Shrub-dominated wetlands Freshwater, tree-dominated wetlands Seasonal/intermittent freshwater marshes/ pools on inorganic soils

Marshes – inorganic soils

Permanent Permanent/seasonal/ intermittent Seasonal/intermittent

Herb dominated Shrub dominated Tree dominated Herb dominated

13. 14.

Nonforested peatlands Forested peatlands

Marshes – peat soils

Permanent

Nonforested Forested

15. 16.

Alpine wetlands Tundra wetlands

Marshes – inorganic or peat soils

High altitude Tundra

17. 18.

Permanent saline/brackish/alkaline lakes Seasonal/intermittent saline/brackish/ alkaline lakes and flats Permanent saline/brackish/alkaline marshes/pools

Lakes

Permanent Seasonal/intermittent

19.

43

20.

Seasonal/intermittent saline/brackish/ alkaline marshes/pools

21.

Geothermal wetlands

22.

Karst and other subterranean hydrological systems

Freshwater

Flowing water

Permanent

Season/intermittent

Permanent Seasonal/intermittent

Saline, brackish, or alkaline water

Rivers, streams, creeks Deltas Springs, oases Rivers, streams, creeks 48 ha o8 ha o8 ha o8 ha

Permanent Marshes and pools Fresh, saline, brackish, or alkaline water

Seasonal/intermittent

Geothermal Subterranean

Derived from Ramsar Convention Secretariat (2006c). Inventory, assessment, and monitoring: An integrated framework for wetland inventory, assessment, and monitoring. In: Ramsar Handbooks for the Wise Use of Wetlands, 3rd edn., vol. 11. Gland, Switzerland: Ramsar Convention Secretariat.

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Managing Aquatic Ecosystems

or accurate classification was required. Semeniuk and Semeniuk (1997) pointed out that the Ramsar classification for inland wetlands was not entirely systematic and promoted the adoption of a more systematic classification based on the hydro-geomorphology of the wetland. Nevertheless, except for the occasional addition of wetland types the Ramsar classification has not been substantially changed and still forms a readily available general model for many purposes.

1.03.3.2 Extent and Distribution Estimates of the global extent of inland aquatic ecosystems differ greatly and highly depend on the definition of wetlands used and on the methods for delineating wetlands (Finlayson et al., 1999; Mitsch and Gosselink, 2007; Whigham, 2009). The Global Review of Wetland Resources and Priorities for Wetland Inventory conducted on behalf of the Ramsar Convention estimated the extent of all wetlands from national inventories as approximately 1280 million hectares, which was considerably higher than previous estimates (Finlayson et al., 1999). This included inland (including lakes, rivers, swamps, and marshes), coastal (including lagoons, swamps, and estuaries), near-shore marine areas (tidal flats and marine areas to a depth of 6 m below low tide), and human-made wetlands (such as reservoirs and rice paddies). Nevertheless, this figure is considered an underestimate, especially for southern America and for certain wetland types (such as intermittently flooded inland wetlands, peatlands, and artificial wetlands) where data were incomplete or not readily accessible. The most recent attempt to ascertain the extent and distribution of inland aquatic ecosystems systems (Lehner and Doll, 2004) based on analysis of existing data sets derived largely from the Earth observation is shown in Figure 7. As with previous estimates, these data contain many inaccuracies and gaps. For example, intermittently inundated inland aquatic ecosystems are not included, and there are many inaccuracies because of the problems of scale and resolution.

Nevertheless, these data and the mapping products are considered incredibly useful even while requiring updating and refinement (Finlayson et al., 2005; Whigham, 2009). Lehner and Doll (2004) reported the extent of inland aquatic ecosystems as 917 million hectares, comprising Africa with 131, Asia 286, Europe 26, Neotropics 159, North America 287, and Oceania with 28 million hectares. Inventory and mapping of inland aquatic ecosystems have been undertaken in many parts of the world, but the level of detail varies from region to region with some regions and ecosystem types considered to be under-represented in the data given above (Finlayson and D’Cruz, 2005). The latter includes rivers, lakes and reservoirs, peatlands, and rice paddies. Inventories of major river systems are available, but there is considerable variability between areal estimates, based on the method and definitions used. Information on the estimated 5–15 million lakes distributed globally is also highly variable and dispersed (WWDR, 2003). Large lakes have been mapped reasonably well, but issues of scale occur with many smaller lakes underestimated or not recorded. Reservoirs are also widespread with the number of large dams (415 m in height) in the world increasing from approximately 5000 in 1950 to more than 45 000 with 3–6 times the standing water held by natural river channels (WCD, 2000; Vo¨ro¨smarty et al., 2005). The overall number of dams globally is uncertain with an estimated 800 000 small dams and further investment in the construction of others, large and small (Vo¨ro¨smarty et al., 2005). The total area of peatlands is estimated as approximately 400 million hectare with the majority in Canada (37%), Russia (30%), USA (13%), and Indonesia (6–7%) (Joosten and Clarke, 2002). The global area of paddies has been estimated to be 130 million hectares (Aselmann and Crutzen, 1989) with almost 90% in Asia. Information on other humanmade wetlands is variable and lacking for some countries. Groundwater systems vary in size from small-scale alluvial sediment along rivers to extensive aquifers such as the 1.2

Lake Reservoir River Freshwater marsh, floodplain Swamp forest, flooded forest Coastal wetland Pan, brackish/saline wetland Bog, fen, mire Intermittent wetland/lake 50−100% wetland 25−50% wetland Wetland complex (0−25% wetland) Figure 7 Distribution of inland aquatic ecosystems described as large lakes, reservoirs, and wetlands. Adapted from Lehner B and Doll P (2004) Development and validation of a global database of lakes, reservoirs and wetlands. Journal of Hydrology 296: 4–22.

Managing Aquatic Ecosystems

million square kilometers of the Guarani aquifer located across parts of Argentina, Brazil, Paraguay, and Uruguay (Danielopol et al., 2003). Groundwater systems have many connections with surface waters although many of these are not well understood, although the karst systems of Slovenia cover nearly 8800 km2 and are well known for their great species biodiversity, while others are not known at all (Finlayson and D’Cruz, 2005; Vo¨ro¨smarty et al., 2005). Whigham (2009) provided a general description of the main features of inland aquatic ecosystems commonly referred to as marshes, peatlands, and swamps. Marshes are dominated by emergent herbaceous vascular plants and occur in areas that are frequently or continuously flooded and most often have mineral soils that do not accumulate peat. Dominant plant species include reeds, rushes, grasses, and sedges, although they can also contain a wide variety of plant species with many different life forms. Peatland is a generic term for inland aquatic ecosystems that have at some point accumulated partially decayed plant matter because of incomplete decomposition, usually to a depth less than 30 cm. Many terms have been used to describe peatlands, such as mires, fens, and bogs. Swamps are flooded intermittently or permanently and are dominated by trees or shrubs. They are diverse and occur from the temperate zones to the tropics with those associated with rivers typically having inorganic substrates while those with little or no connection to flowing streams may develop peat substrates. While these generic descriptions exist, there are many local variations and terms used to describe inland aquatic ecosystems.

1.03.3.3 Loss and Degradation of Inland Aquatic Ecosystems The loss and degradation of inland aquatic ecosystems have been reported from many parts of the world (Finlayson et al., 1992; Mitsch, 1998; Moser et al., 1996; Whigham, 2009), but there are few reliable estimates of the actual extent of this loss globally. Dugan (1993) speculated that about 50% of wetlands had been lost globally, but did not provide supporting evidence. As a reliable estimate of the extent of inland aquatic ecosystems, particularly peatlands and intermittently inundated wetlands in semiarid areas, is not available (Finlayson et al., 1999), it is not possible to ascertain the extent of wetland loss globally (Finlayson and D’Cruz, 2005). While there is considerable uncertainty surrounding the extent of wetland losses globally, there is a lot of information available from specific countries or regions. Some examples are given below, taken from a collation provided in the Millennium Ecosystem Assessment (Finlayson and D’Cruz, 2005). The information available on the loss of inland aquatic ecosystems is far better for North America than for many other parts of the world with systematic monitoring of wetlands, excluding lakes and rivers, in the United States showing a loss of 116 000 ha yr 1 from the mid-1970s to the mid-1980s decreasing to 23 700 ha yr 1 from 1986 to 1997 (Dahl and Johnson, 1991; Dahl, 2000). Most of this loss was from the conversion or drainage of wetlands for urban development and agricultural purposes with an estimated 42.7 million hectares of wetlands remaining out of the 89 million hectares estimated to have been present at the time of European

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colonization (Dahl, 2000). The decline in the rate of loss was attributed largely to the successful implementation of wetland policies and programs that promoted the restoration, creation, and enhancement of wetlands, as well as incentives to deter draining of wetlands. This is shown by a net gain of about 72 870 ha of wetland after 1997 and a 47 000 ha increase in the area of lakes and reservoirs (Dahl, 2000). While the data are not as illustrative for other parts of the world, Finlayson and D’Cruz (2005) concluded from the literature that much of the loss of wetlands in the northern temperate zone occurred during the first half of the twentieth century. Since the 1950s, many tropical and subtropical wetlands, particularly swamp forests, have also been lost or degraded, particularly as a consequence of agricultural expansion. The OECD (1996) estimated that by 1985, 56– 65% of available wetland had been drained for intensive agriculture in Europe and North America, 27% in Asia, 6% in South America, and 2% in Africa – a total of 26% loss to agriculture worldwide. This is still occurring, for example, in South America where peatlands linked with the Andean paramos ecosystems are being converted for agriculture, forestry, and peat mining (Hofstede et al., 2003; Blanco and de la Balze, 2004). In Southeast Asia large areas of the once-extensive tropical peat swamp forests have been degraded or lost over the last four decades mainly because of logging for timber and pulp and more recently by clear-felling and conversion to oil palm plantations (Glover and Jessup, 1999; Page et al., 1997; Rieley and Page, 1997). The most dramatic loss of peatlands to agriculture has been in northern Europe in countries, including Finland, Estonia, Denmark, the United Kingdom, and the Netherlands (once one-third peatland) which has lost virtually all of its natural peatlands (Brag et al., 2003; Joosten, 1994). Despite the absence of national data, there are many welldocumented examples of large inland aquatic ecosystems that have been degraded or lost. These include the impacts of water diversions to the Aral Sea in Central Asia (Lemly et al., 2000) and the Mesopotamian marshes in Iraq (Richardson et al., 2005), the Murray-Darling Basin in Australia (Kingsford and Johnson, 1998), the Everglades in the United States (Richardson, 2008), Donana in Spain (Bartolome and Vega, 2002), and the Hadejia-Nguru wetland complex in Nigeria (Lemly et al., 2000). The Aral Sea in Central Asia represents one of the most extreme cases in which water diversion for irrigated agriculture has caused severe environmental degradation of an inland water system with detrimental impacts on human well-being (see summary in Finlayson and D’Cruz, 2005; Box 2 and Figure 8). The extent of adverse change is so severe that the Aral Sea is considered by Falkenmark et al. (2007) as an example of human modification of an inland aquatic system having gone too far.

1.03.3.4 Loss of Species from Inland Aquatic Ecosystems The extent of species loss from inland aquatic ecosystems has been documented through several programs in recent years (Revenga et al., 2000; Finlayson and D’Cruz, 2005; Dudgeon et al., 2005; Loh et al., 2005). Information from these documents has been used to summarize the extent of species losses. The following excerpt from Revenga et al. (2005) serves as an

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Managing Aquatic Ecosystems

Box 2 The Aral Sea. Based on information derived from Finlayson CM and D’Cruz R (2005) Inland Water Systems Millennium Ecosystem Assessment, Volume 2: Conditions and Trends. Washington, DC: Island Press and Falkenmark M, Finlayson CM, and Gordon L (2007) Agriculture, water, and ecosystems: Avoiding the costs of going too far. In: Molden D (ed.) Water for Food, Water for Life: A Comprehensive Assessment of Water Management in Agriculture, pp. 234–277. London: Earthscan. The Aral Sea is one of the most prominent examples of how unsustainable water management has led to a large-scale and possibly irreversible ecological and human disaster. Drastically reduced water flow into the sea has impaired human livelihoods and health, affected the local climate, and reduced if not decimated much of the biodiversity. Since 1960 the volume of water in the basin that surrounds the Aral Sea Basin has been reduced by 75% mainly as a consequence of the development of almost 7 million hectares of irrigation (UNESCO, 2000; Postel, 1999). It is considered unlikely that the ecological and social changes that have occurred as a consequence will be successfully restored, despite efforts to rehabilitate the northern part of the sea.

1973

1986

1999

2001

Figure 8 Changes in the Aral Sea 1973–2001. Reproduced from UNEP (2005).

introduction to the loss of species from inland aquatic ecosystems, including mention of the drivers, the relative risk of extinction compared to other ecosystems, and the inadequate level of assessment and information: Human activities have severely affected the condition of freshwater ecosystems worldwide. Physical alteration, habitat loss, water withdrawal, pollution, overexploitation and the introduction of nonnative species all contribute to the decline in freshwater species. Today, freshwater species are, in general, at higher risk of extinction than those in forests, grasslands and coastal ecosystems. For North America alone, the projected extinction rate for freshwater fauna is five times greater than that for terrestrial fauna – a rate comparable to the species loss in tropical rainforest. Because many of these extinctions go unseen, the level of assessment and knowledge of the status and trends of freshwater species are still very poor, with species going extinct before they are even taxonomically classified.

As with inland aquatic ecosystems the data on the condition and trends of freshwater species are, for the most part,

poor at the global level, although some countries have reasonable inventories. Key conclusions from Revenga and Kura’s (2003) assessment of the level of knowledge of the distribution and condition of inland water biodiversity were: fish and waterbirds were by far the best-studied groups, although with considerable regional differences; aquatic plants, insects, freshwater mollusks, and crustaceans were poorly known in most parts of the world, with fragmentary information; and that every group of organisms considered, including aquatic plants, invertebrate, and vertebrate animal species, contained examples of extinct, critically endangered, endangered, and vulnerable taxa. This contrasts with the importance of inland aquatic ecosystems which were reported by McAllister et al. (1997) to be species rich relative to other ecosystems and to support a disproportionately large number of species of some taxonomic groups, for instance, some 40% of known species of fish and about 25–30% of all vertebrate species diversity (Leveque et al., 2005). The living planet index developed by WWF and UNEPWCMC (Loh and Wackernagel, 2004) provides a measure of the trends in more than 3000 populations of 1145 vertebrate species around the world. The 2004 freshwater species population index, which took the trend data into account for 269 temperate and 54 tropical freshwater species populations (93 of which were fish, 67 amphibians, 16 reptiles, 136 birds, and 11 mammals), showed that freshwater populations declined consistently and at a faster rate than the other species groups assessed, with an average decline of 50% between 1970 and 2000 (Figure 9). Over the same period, both terrestrial and marine fauna decreased by 30% (Figure 9). A summary of the status of separate groups of species is given in Box 3. Revenga et al. (2005) reported that a review by the World Resources Institute of the status and trends of inland water biodiversity for the Convention on Biological Diversity (Revenga and Kura, 2003) drew the following conclusions:

1. freshwater fishes and waterbirds were by far the beststudied groups of species from inland aquatic ecosystems, although there were considerable regional differences; 2. aquatic plants, insects, freshwater mollusks, and crustaceans were poorly known or assessed in most parts of the world, with only fragmentary information available; and 3. in every group of organisms considered there were examples of extinct, critically endangered, endangered, and vulnerable taxa, making it clear that inland aquatic ecosystems were among the most threatened of all environments.

Managing Aquatic Ecosystems

120

100

Index 100 in 1970

Terrestrial 80 Marine 60 Freshwater 40

20

0 1970

1975

1980

1985

1990

1995

2000

Figure 9 Trends in freshwater, marine, and terrestrial living planet indices, 1970–2000. From Finlayson CM, D’Cruz R, and Davidson NJ (2005) Ecosystem Services and Human Well-Being: Water and Wetlands Synthesis. Washington, DC: World Resources Institute; and Millennium Ecosystem Assessment (2005) Ecosystems and Human Well-Being: Synthesis. Washington, DC: Island Press.

They further concluded that in general the information on species of inland aquatic ecosystems was poor, even for economically important groups, such as fish, and pointed out that many inventories tended to be organized by taxonomic groups and not by ecosystem types which made it hard to assess the condition of aquatic ecosystems. These conclusions have been largely supported by the analyses undertaken in the Millennium Ecosystem Assessment (Finlayson and D’Cruz, 2005; Finlayson et al., 2005) and more generally by other assessments, such as the Global Environment Outlook (Arthurton et al., 2007).

1.03.4 Drivers of Change in Inland Aquatic Ecosystems Analyses over the past two decades have identified a suite of common drivers of change in inland aquatic ecosystems (e.g., Revenga and Kura, 2003; Whigham et al., 1993; Mitsch, 1994; Finlayson and D’Cruz, 2005; Dudgeon et al., 2005). Many of these previous reviews have focused primarily on biophysical pressures that directly affect the ecological condition of these ecosystems, as depicted in Figure 10. Also, the importance of addressing the indirect drivers of change has been increasingly recognized with the Millennium Ecosystem Assessment (2003) and the Global Environment Outlook (Arthurton et al., 2007) providing comprehensive overviews. The indirect drivers are derived primarily from the following: demographics such as population size, age, gender structure, spatial distribution, and migration; economic forces such as national and per capita income, macroeconomic policies, international trade, and capital flows; sociopolitical

47

processes such as governance, institutional and policy frameworks, and the roles of women and wider civil society; scientific and technological developments including rates of investments in research and development and the rates of adoption of new biotechnologies and information technologies; and cultural and religious choices individuals make about what and how much to consume and what they value. These drivers are not static – they can change rapidly and over long periods as shown by fluctuations in the global economy and the increasing global population, for example. Changes in these drivers are also expected to increase the demand for food, fiber, energy, and freshwater (Vo¨ro¨smarty et al., 2005; Molden et al., 2007). The direct drivers of change in inland aquatic ecosystems are interconnected with the indirect drivers, and include: changes in land use as a consequence of clearance, drainage, and infilling; the spread of infrastructure for urban, tourism and recreation, aquaculture, agriculture, and industrial purposes; the introduction and spread of invasive species; the regulation and fragmentation of rivers; abstraction of surfaceand groundwater; over fishing and in places unsustainable hunting; chemical pollution, salinization, and eutrophication; and more recently the impacts of global climate change. In some cases, these drivers act synergistically or cumulatively. Multiple interacting drivers can cause changes in aquatic ecosystems and their species and ecosystem services that may not be readily attributable to one or the other driver. There are many interdependencies between and among the indirect and direct drivers of change, and, in turn, changes in ecosystems can lead to feedbacks on the drivers of change. When addressing the complex scenarios of multiple drivers, it is necessary to have a clear understanding of the nature of the changes and their likely causes before implementing management responses; risk and vulnerability assessments can help ascertain the nature of change and guide management. The Millennium Ecosystem Assessment (2003) outlined the interactions between drivers and ecosystems in a framework that linked the consequences of indirect and direct drivers with changes in the biodiversity and services provided by ecosystems and the consequences for human well-being. The direct drivers of change in inland aquatic ecosystems are described below – it draws heavily on the assessment undertaken by Finlayson and D’Cruz (2005) as part of the Millennium Ecosystem Assessment and provides an update.

1.03.4.1 Drainage, Clearing, and Infilling It has been well established that clearing or drainage for agricultural expansion is the principal cause for wetland loss worldwide. Agriculture, including rangelands, now occupies roughly 40% of the world’s terrestrial surface and is a major contributor to global environmental change (Foley et al., 2005), with cropping occurring on more than 50% of the land area in many river basins in Europe and India and more than 30% in the Americas, Europe, and Asia (Millennium Ecosystem Assessment, 2005). In this respect, inland aquatic ecosystems are subject to many of the same pressures from agriculture as are other ecosystems – agriculture is recognized as a major driver of change in ecosystems globally (Millennium Ecosystem Assessment, 2005).

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Managing Aquatic Ecosystems

Box 3 Status of animals from inland aquatic ecosystems. Based on information derived from Finlayson CM and D’Cruz R (2005) Inland water systems Millennium Ecosystem Assessment, Volume 2: Conditions and Trends. Washington, DC: Island Press. Invertebrates The conservation status of most aquatic invertebrates has not been comprehensively assessed, except for regional assessments of certain taxonomic groups. Clausnitzer and Jodicke (2004) assessed the global status of dragonflies and damselflies and reported that while some 130 species were previously listed as threatened that many more were now also threatened. IUCN (2003) reported that 130 freshwater species of aquatic insects, 275 species of freshwater crustacean, and 420 freshwater mollusks were globally threatened, although no comprehensive global assessment has been made of all the species in these groups. For the United States, one of the few countries to assess freshwater mollusks and crustaceans comprehensively, 50% of known crayfish species and two-thirds of freshwater mollusks are at risk of extinction, and at least one in 10 freshwater mollusks are likely to have already gone extinct (Master et al., 1998). Freshwater fish A number of regional overviews of the status of freshwater fish are available, yet many of the existing overviews underestimate the number of species, as there are still many species to be described and assessed. There is, therefore, a high level of uncertainty about the status of fish in many inland waters with estimates of the number of freshwater fish in Latin America varying from 5000 to 8000; in tropical Asia and Africa, there are estimated 3000 species on each continent (Revenga and Kura, 2003), although these figures are almost certainly underestimates. It is estimated that in the last few decades more than 20% of the world’s 10 000 described freshwater fish species have become threatened or endangered or are listed as extinct (Moyle and Leidy, 1992). In the 20 countries for which assessments are most complete, an average of 17% of freshwater fish species are globally threatened (IUCN, 2003). Amphibians The recent Global Amphibian Assessment (IUCN et al., 2004) reported that the decline in conservation status of freshwater amphibians was worse than that of terrestrial species listed with 964 of 3908 freshwater species listed as threatened. Species associated with flowing water were found to have a higher risk of extinction than those associated with still water. Salamanders and newts have an even high level of threat (46% globally threatened or extinct) than frogs and toads (33%) and Caecilians (2%, although knowledge of these is poor, with only one-third assessed). Basins with the highest number of threatened freshwater amphibians include the Amazon, Yangtze, Niger, Parana, Mekong, Red, and Pearl in China, Krishna in India, and Balsas and Usumacinta in Central America, all of which have between 13 and 98 threatened freshwater species. Reptiles Van Dijk et al. (2000) reported that of the 200 species of freshwater turtles, 51% of the species of known status have been assessed as globally threatened, and the number of critically endangered freshwater turtles more than doubled in the four years preceding and that of the 90 species of Asian freshwater turtles and tortoises, 74% are considered globally threatened. Of the 23 species of crocodilians, which inhabit a range of wetlands including marshes, swamps, rivers, lagoons, and estuaries, four are critically endangered, three are endangered, and three are vulnerable (IUCN, 2003). There is very little information on the conservation status of aquatic snakes although IUCN (2003) reported that some semiaquatic snakes are vulnerable. Waterbirds Many waterbird species are globally threatening and the status of both inland and marine/coastal waterbirds is deteriorating faster than those in other habitats (Davidson and Stroud, 2004). Of the 35 bird families with species that are entirely or predominantly coastal/marine or inland wetland dependent, 20% of the 1058 species for which assessment data exist are currently globally threatened or extinct. Waterbirds dependent on freshwater ecosystems, especially those that also use marine and coastal ecosystems, have deteriorated in status faster than the average for all threatened species, but at similar rates for other migratory bird species. Shorebirds, many of which also use freshwater ecosystems, are declining worldwide with 48% of populations with a known trend declining. Other waterbirds in decline include cranes with 47% of populations with a known trend declining, rails (50%), skimmers (60%), darters (71%), ibis and spoonbills (48%), storks (59%), and jacanas (50%). Only gulls (18%), flamingos (18%), and cormorants (20%) appear to have a relatively healthy status. Mammals Although most mammals depend on freshwater for their survival, and many feed in rivers and lakes or live in close proximity to freshwater ecosystems, only a few are considered aquatic or semiaquatic mammals. Revenga and Kura (2003) provided an analysis of the status of aquatic and semiaquatic mammals. Some 37% of inland water-dependent mammals are globally threatened, compared with 23% of all mammals and includes otters (50% of species of known status threatened), seals (67% threatened), manatees (100% threatened), river dolphins and porpoises (100% threatened), and wetland-dependent antelopes (29% threatened) (Revenga et al., 2005).

Peatlands in particular have for centuries been converted for agriculture in many parts of the world, particularly in Europe, but also more recently in the highlands of South America and in parts of China, Southeast Asia, and Africa. In Southeast Asia, large areas of the once-extensive tropical peat swamp forests have been heavily degraded, and large extents have been lost over the last four decades as a consequence of logging for timber and pulp and conversion to oil palm plantations. The peatlands of Malaysia and Indonesia are especially threatened by drainage and forest clearing that then makes them susceptible to fire. Land clearing and subsequent uncontrolled fires in 1997 severely burned about 5 million

hectares of forest and agricultural land on the Indonesian island of Borneo (Page et al., 2002, 2009). Irrigated agriculture is another major driver of the loss and degradation of inland aquatic ecosystems with water withdrawals for irrigation worldwide resulting in major changes in river flows (Revenga et al., 2000) – flows that are essential for sustaining the ecosystem services and species that occur in inland aquatic ecosystems. The global extent of irrigated agricultural land has increased from approximately 138 million hectares in 1961 to 271 million hectares in 2000, and currently accounts for an estimated 40% of total food production even though it represents only 17% of global

Managing Aquatic Ecosystems

49

Dams

Overharvesting of wild resources,

interrupt the connectivity of river systems, disrupting fish spawning and migration. Dams with large reservoirs alter seasonal flood regimes and retain sediment needed to maintain the productivity of floodplain agriculture.

especially fish, is driven both by the subsistence needs of a growing population and by unsustainable commercial exploitation, threatening future food security and livelihoods.

River channelization and dredging for navigation reduces riverine habitat and alters flood patterns.

Large-scale irrigation and river diversions after natural flow regimes, reduce downstream water availablity for agriculture, and contribute to salinization through saltwater intrusion in the coastal zone.

Agricultural expansion is often achieved by converting natural inland water systems, reducing aquatic biodiversity and natural flood control functions, and increasing soil salinity through evaporation. When accompained by intensive use of agrochemicals, off-site pollution effects can be extensive.

Forest clearing in permanently or seasonally inundated zones, often motivated by unsustainable aquaculture production, dramatically reduces habitat for wild aquatic organisms. In the coastal zone, it also make the landscape much more susceptible to erosion.

Roads and flood-control infrastructure often interrupt wetland connectivity, disrupting aquatic habitat, reducing the function of wetlands to remove pollutants and absorb floodwaters, and potentially increasing the losses when high floods do occur.

Urban and industrial pollution, when released untreated into aquatic environments, reduces water quality, affecting the diversity and abundance of aquatic organisms as well as human health.

Figure 10 Pictorial representation of some of the direct drivers of change in inland and coastal aquatic ecosystems. Invasive species, climate change, and land conversion to urban or suburban areas affect all components of the catchment and coastal zone and are therefore not represented pictorially. From Millennium Ecosystem Assessment; adapted from Ratner BD, Ha DT, Kosal M, Nissapa A, and Chanphengxay S (2004) Undervalued and Overlooked: Sustaining Rural Livelihoods through better Governance of Wetlands, Studies and Review Series. Penang, Malaysia: World Fish Centre.

cropland area (Wiseman et al., 2003). In this respect, it seems to have a disproportionate negative impact on inland aquatic ecosystems relative to the land area involved. Possibly more significantly, around 66% of all water withdrawn for direct human use is now being used for agriculture (Scanlon et al., 2007). The problem of increasing abstraction of freshwater from aquatic ecosystems is exacerbated by the loss of much of this water from the immediate landscape – very little of the water returns as runoff to the rivers as most of it is evaporated or transpired (Falkenmark and Lannerstad, 2005). There are many well-documented examples where diversion of water for agriculture has caused a decline in the extent and degradation of inland aquatic ecosystems and their species richness (Revenga et al., 2000; Finlayson and D’Cruz, 2005; Dudgeon et al., 2005). Lake Chad provides an interesting example with major ecosystem change being due to

both human-induced and natural changes, with the loss of many species and ecosystem services as the lake shrank from about 25 000 km2 in surface area to one-twentieth of its size over 35 years at the end of the twentieth century as a consequence of a drier climate and high agricultural demands for water (Coe and Foley, 2001). This example illustrates the complexity that can arise when multiple drivers of change impact on an aquatic ecosystem, especially in areas of high climate variability.

1.03.4.2 Modification of Water Regimes Agricultural development and the diversion of water for irrigation, and increasingly for urban purposes, have modified the water regime in many inland aquatic ecosystems. Modifications include the construction of river embankments to

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Table 4

Alteration of inland freshwater systems worldwide

Alteration

Pre-1990

1900

Waterways altered for navigation (km) Canals (km) Large reservoira Number Volume (km2) Large dams (415 m high) Installed hydro-capacity (MW) Hydro-capacity under construction (MW) Water withdrawals (km3 yr 1) Wetlands drainageb (km3)

3125 8750

8750 21 250

41 14 – – – – –

581 533 – – – 578 –

1950–60 – – 1 105 1 686 5 749 o290 000 – 1 984 –

1985 4500 000 63 125 2768 5879 – 542 000 – 3200 160 000

1996–98

– 2836 6385 41 413 660 000 126 000 3800 –

a

Large reservoirs are those with a total volume of 0.1 km3 or more. This is only a subset of the world’s reservoirs. Includes available information for drainage of natural bogs and low-lying grasslands as well as disposal of excess water from irrigated fields. – Data not available. From Revenga C and Kura Y (2003) Status and Trends of Biodiversity of Inland Water Ecosystems, Technical Series No. 11. Montreal: Secretariat of the Convention on Biological Diversity.

b

improve navigation, drainage of wetlands for agriculture, construction of dams and irrigation channels, and the establishment of interbasin connections and water transfers. Revenga and Kura (2003) provided data on the extent of alteration to inland freshwater systems worldwide (Table 4). These changes have had many beneficial outcomes for people through the provision of local flood control and hydropower, improved fisheries and increased agricultural output (Molden et al., 2007), but at the same time there have been many negative ecological effects on inland aquatic ecosystems (Revenga et al., 2000). Rivers have been disconnected from their floodplains and wetlands; seasonal changes in water flows have disrupted fish and bird migration and breeding; greater runoff in rivers has increased the likelihood and severity of flooding; and links with groundwater systems have been disrupted, and, in some coastal regions, enabled saline water to intrude on freshwater systems. They have also transformed many rivers through (1) the construction of large reservoirs, such as those on the Volta and Zambezi Rivers in Africa, or along the Volga River in Russia; (2) the embankment and channelization of rivers such as that along the Mississippi and Missouri rivers in the United States; or (3) significantly reduced flows to floodplains and downstream ecosystems, including deltas such as the Indus in Pakistan, or the lakes at the mouth of the Murray River in Australia. Even the large inland seas are not safe from the impacts of river regulation and diversion of water away from terminal water bodies. The Dead Sea located in the Syrina-African rift valley at the southern outlet of the Jordan River and at 417 m below sea level is the world’s saltiest large water body. It is threatened by excessive withdrawal of water from the river to support industrial, agricultural, and tourism development (ILEC and UNEP, 2003). The annual historical flow of the Jordan River to the Sea was about 1285 million cubic meters in the 1950s compared to 505 in mid-1970s, 275 in 2000s, and a projected 170 million cubic meters in the mid-2020s (Courcier et al., 2005). In addition to the reduction in water flows, the diversion of water has resulted in the development of a complex socioeconomic system (Figure 11) with undoubted economic and political ramifications associated with any proposals for

further development and regulation of water flows, including management of wastewater and irrigated agriculture (Courcier et al., 2005). The Mesopotamian marshlands also provide another example of a complex social–political scenario associated with river regulation and the restoration of an intermixed social and ecological system. The marshes have been severely affected by river regulation in recent decades with the original areas of 15 000–20 000 km2 before being reduced by drainage and dam construction along the Tigris and Euphrates rivers to less than 400 km2 (Partow, 2001). The total capacity of the reservoirs along these rivers exceeds the annual discharge of both rivers, drastically reducing the supply of flood waters that were so important for delivering sediments and nutrients to the marshlands. In addition to regulating the flows along the rivers by the construction of dams in the upstream reaches, attempts were made in the early 1990s to drain water away from the marshes through large canals. The combined effect of these moves was to reduce the extent of the marshes and threatened the culture and biodiversity that depended on the annual flooding regime. More recently, there have been partially successful but still insufficient efforts to restore parts of the marshes by breaking banks and flooding some 20% of the original area of marsh (Richardson et al., 2005). While these efforts have indicated the potential for further successful restoration, Falkenmark et al. (2007) have cautioned that attempts to return water to the central areas of the marshes upstream of the confluence of the Tigris and Euphrates could generate adverse impacts on aquatic ecosystems further downstream. That is, without an increase in the amount of water available for flooding the marshes, simply returning the water to upstream areas may not be enough to restore them and could further reduce the flow of water to downstream areas and possibly further reduce the flow to the Persian Gulf. Richardson et al. (2005) have also reported the construction of a dike along the Iraq/Iran border that could further reduce the flow of water into the marshes. The construction of large dams has doubled or tripled the residence time of river water with impacts on suspended sediment and carbon fluxes, waste processing, and aquatic

Managing Aquatic Ecosystems

1950

2000s

51

Mid-1970s

Mid-2000s

Figure 11 Changes in water resources and their allocation in the lower Jordan basin from the 1950s to the mid-2020s. Reproduced from Courcier R, Venot J-P, and Molle F (2005) Historical Transformations of the Lower Jordan River Basin (in Jordan): Changes in Water Use and Projections (1950– 2025). Colombo, Sri Lanka: International Water Management Institute.

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habitat, and has resulted in fragmentation of the river channels with 37% of 227 river basins around the world strongly affected by fragmentation and altered flows, 23% moderately affected and 40% unaffected (Revenga et al., 2000). Small dams can also have major effects on the ecological condition of inland aquatic ecosystems. The debate about the construction of dams is ongoing (WCD, 2000). The effects of modification of flow regimes on fish migrations have been reviewed by Revenga and Kura (2003) with direct impacts on diadromous fish species such as salmon being well known and increasingly recognized, whereas the indirect impacts of flow alteration, such as the reduction of floods and loss of lateral connections on floodplains, are not always as evident. In many cases, construction of dams has resulted in the disappearance of fish species adapted to river systems and the proliferation of species adapted to lakes. Changes in the fish are indicative of many changes in the biodiversity of regulated rivers and associated aquatic ecosystems, although the extent of data and information about the wider biodiversity is often inadequate or lacking (Revenga et al., 2005).

1.03.4.3 Invasive Species Despite the current concern about invasive species in inland aquatic ecosystems their importance has not always been as widely appreciated (Finlayson, 2009). This was in part because the problem of invasive species in these ecosystems was seen largely as one for developed countries, despite the paradoxical occurrence of well-documented cases of invasive species in African wetlands and lakes (e.g., Salvinia molesta and Nile perch – Lates nilotica). The reasons for this situation are not clear, although they probably included insufficient awareness and information about the impacts of these species and ways of controlling them. However, several recent assessments (such as the Millennium Ecosystem Assessment (2005)) and initiatives (such as the Global Invasive Species Programme) have demonstrated that the loss of wetland species as a consequence of invasion by invasive species is now much more of a concern globally (Revenga et al., 2005; Finlayson, 2009). The spread and establishment of non-native invasive species in inland aquatic ecosystem have caused many changes to the native biota and are likely to become even more common

Box 4

with the further development of aquaculture, interbasin transfers of water, and shipping and global commerce. There are many documented and well-known examples of plant species that have successfully invaded and established in inland aquatic ecosystems, including the pan-tropical weeds salvinia (Salvinia molesta) and water hyacinth (Eichhornia crassipes) that originated in South America but are now widely established in many countries. In many instances though, the occurrence of alien plant species may not be seen as undesirable, as shown by the establishment of the alien species Egeria densa in the Rio Cruces wetland in Chile where it was considered to be an ecological engineer and thrived and provided the mainstay to support a population of blacknecked swans (Cygnus melancoryphus) which was highly appreciated and valued by local residents (Yarrow et al., 2009). The angst in this case came not from the establishment of the alien invasive plant but from its decline when its biomass crashed suddenly in 2004 and the swans dispersed to other wetlands, leaving an acrimonious debate about the cause of the population crash (Delgado et al., 2009). The example of Canadian pond weed (Elodea canadensis) outlines many of the dilemmas raised by introduced species (Sculthorpe, 1967). It originated in North America and invaded the waterways of Europe in the late nineteenth century where it grew rapidly and spread vegetatively to reach a maximum population density within a period of a few months to 4 years. This population level was maintained for up to 5 years but then declined to levels that were not considered a nuisance. The reasons for the rapid increase and decline were not determined. Many animal species, both large and microscopic, have also invaded inland aquatic ecosystems, such as those outlined for European lakes (Box 4). The larger invasive animals include the cane toad (Bufo marinus), bullfrog (Rana catesbeiana), European domestic pig (Sus scrofa), carp (Cyprinus carpio), and zebra mussel (Dreissena polymorpha) that have become established outside of their native range and disrupted the inland water systems that they have invaded. Many fish species have been spread beyond their native ranges often in response to demands for aquaculture and aquarium species. Fish introductions have usually been done to enhance food production and recreational fisheries or to control pests such as mosquitoes and aquatic weeds. The spread of trout and salmon

Invasive species in European rivers. Based on information supplied by H. Ketelaars.

The construction of canals between rivers and other water bodies in Europe over the past two centuries has provided channels for the migration of many aquatic species, whether they migrated themselves or were carried by on the hulls of ships or in their ballast water. The Volga–Baltic Waterway, reconstructed in 1964 to connect the Caspian with the Baltic enabled the translocation of many aquatic species, including copepods, rotifers, the onychopod Bythotrephes longimanus, and several fish species to the Volga basin. The Main-Danube Canal, officially opened in 1992, is another that allowed many Ponto-Caspian invertebrate species to reach the Rhine basin and from there to disperse to other basins, mainly in ballast water. Intentional introductions of aquatic species have occurred mainly in the past two centuries. The North American amphipod Gammarus tigrinus was deliberately introduced in 1957 to Werra and Weser rivers in Germany where the local gammarid fauna had disappeared due to excessive chloride pollution. The mysid Mysis relicta was been introduced to many Scandinavian lakes to stimulate fish production. Three North American introduced crayfish species have established themselves in many European waters and with the introduced crayfish plague (Aphanomyces astaci) have almost eliminated the native crayfish (Astacus astacus). At least 76 freshwater fish species have been introduced into European fresh waters, with approximately 50 establishing self-sustained populations. When introductions between areas within Europe are also considered, the number of introduced fish species is more than 100. The numerically most important families are cyprinids and salmonids, of which grass carp (Ctenopharyngodon idella), silver carp (Hypophthalmichthys molitrix), rainbow trout (Oncorhynchus mykiss), and brook char (Salvelinus fontinalis) are now widely distributed.

Managing Aquatic Ecosystems

species for sports fishing is well known. The introduction of alien fish species has though often resulted in major ecological change, including the collapse of native fish populations. Finlayson and D’Cruz (2005) provided a summary of impacts of some invasive species, including the adverse impact of salmonids on the genetic diversity of wild stocks in many countries and the spread of tilapia species into Central and Southern America and parts of Asia. Herbivorous and omnivorous species, such as Indian, Chinese, and common carp, account for the majority of introductions in tropical Asia (Revenga and Kura, 2003). In many cases, the impact of invasive species on the native fish has not been documented. Finlayson (2009) noted that while there are some obvious examples of species that have established outside their native range, it should not be assumed that all newly established species have been transplanted by human activities or are invasive. Recent concerns over global climate change and variability provide a scenario where it may no longer be possible to attribute the occurrence of new species to natural fluctuations versus human activity.

1.03.4.4 Overfishing Inland fisheries are a major source of protein for a large proportion of the world’s population with the global production of fish and fishery products from inland waters in 2002, amounting to 32.6 million tonnes with 8.7 tonnes from wild capture and the rest from aquaculture (FAO, 2004). While inland fisheries have increased, FAO (1999) also reported that most inland fisheries that relied on natural reproduction of fish stock were overfished or being fished at their biological limit. Arthurton et al. (2007) also reported that inland fish stocks were subjected to a combination of direct pressures, including habitat alteration, and loss, altered flows, and habitat fragmentation due to dams and other infrastructure, and also faced problems from pollution, exotic species, and overfishing. With much of inland fisheries catches destined for subsistence consumption or local markets, food demand for growing populations is a major factor driving exploitation levels in inland waters. Monitoring of the extent of fishing in inland aquatic systems also seems to be underreported by a factor of 2 or 3, due to the large volume of harvest that is consumed locally, and remains unrecorded (Allan et al., 2005). Even with discrepancies in the data, it is well established that inland fisheries are extremely important in Asia and Africa and in 2002 accounted for 90% of the inland fish catch (FAO, 2004). China alone accounted for at least one-quarter of the inland catch, followed by India (9%), Bangladesh (8%), and Cambodia (4%) (FAO, 2004). The importance of aquaculture as a component of inland fish supply is shown by continued growth at an average rate of nearly 9% per year since 1970 – a much higher rate than that for capture fisheries (B1%) (FAO, 2004). Almost 58% of this came from China with an average annual increase of 11% between 1970 and 2000, compared with 7% for the rest of the world (FAO, 2004). However, many aquaculture operations, depending on their design and management, can contribute and have contributed to habitat degradation, pollution,

53

introduction of exotic species, and the spread of diseases through the introduction of pathogens (Naylor et al., 2000). There is ample evidence that overfishing is a significant factor in the decline of numerous species and fisheries, and is of global importance as a threat to inland water biodiversity (Allan et al., 2005). There are two main types of overfishing with intensive fishing of a targeted species leading to marked declines in catch per unit effort and size of individuals captured, while assemblage or ecosystem overfishing leads to sequential declines of species and depletion of individuals and species of large size, especially piscivores, and declines in the mean trophic level of the assemblage and changes in the responsiveness of populations to environmental fluctuations. The historic influence of overharvesting of fish is shown by the decline of the Murray cod of the Murray-Darling river system in Australia, some sturgeon stocks of Eurasia, the tilapiine species Oreochromis esculentus and Oreochromis variabilis of Lake Victoria, and perhaps the Pacific salmon of the Columbia River while the decline of the Mekong giant catfish and the Nile perch of Lake Victoria provide contemporary cases (Allan et al., 2005; Box 5). The consequences of eliminating fish species from inland aquatic ecosystems are likely to be numerous and of varying severity with the progressive reduction in assemblage diversity meaning that fewer species are available to perform critical functions in the ecosystem with dire consequences following the loss of a species with a disproportionately strong influence on nutrient, habitat, or assemblage dynamics (Allan et al., 2005).

1.03.4.5 Water Pollution and Eutrophication Vo¨ro¨smarty et al. (2005) noted that attempts to summarize patterns and trends in the quality of inland waters, particularly at a global scale, encompassed an array of challenges that included basic definitional problems, a lack of worldwide monitoring capacity, and an inherent complexity in the chemistry of both natural and anthropogenic pollutants. Furthermore, despite improvements in analytical methods the capacity to monitor trends in water quality is limited in terms of the spatial coverage, frequency, and duration of monitoring data. Data comparability was yet another constraint while the monitoring of groundwater was more problematic than surface water. The Millennium Ecosystem Assessment (2005) highlighted changes in the global nitrogen cycle with the loading of reactive nitrogen to the landmass having doubled from 111 million to 223 million tonnes per year with the greatest increases in North America, continental Europe, and South and East Asia (Green et al., 2004). This has caused the transport of dissolved inorganic nitrogen in rivers to increase from about 2–3 million tonnes per year from preindustrial times to about 15 million tonnes today, especially in drainage basins that are heavily populated or supporting extensive industrial agriculture. While rivers and wetlands can assimilate some nitrogen, the self-purification capacity is not unlimited and the water quality in many has deteriorated, resulting in eutrophication, harmful algal blooms, and high levels of nitrate in drinking water (Malmqvist and Rundle, 2002). Jorgensen et al. (2001) reported that eutrophication was a widespread problem in lakes and reservoirs and also one of the most

54

Managing Aquatic Ecosystems

Box 5 Nile Perch invasion in Lake Victoria, East Africa. Based on information from Howard (2009) Case study 1 Nile Perch invasion in Lake Victoria. In: Finlayson CM (2009) Biotic pressures and their effect on wetland functioning. In: Maltby E and Barker T (eds.) The Wetlands Handbook, pp. 674–676. Oxford: Wiley-Blackwell. The Nile perch (Lates niloticus) is a large predatory fish that can grow to 1.8 m in length and weigh as much as 200 kg. It is native to the White Nile River system in Uganda, Sudan, Ethiopia, and Egypt and was introduced into Lake Victoria in the 1950s and 1960s. At the time the lake contained around 350 species, including 300 endemic cichlids of the subfamily Haplochromiinae that occupied many niches in the lake and its wetlands. The Nile perch was introduced to Lake Victoria to boost its fishery, but was hardly seen in catches until the late 1970s with the entire lake yielding less than 25 000 tonnes in 1981 but rising to 363 000 tonnes in 1993. At the same time, the total catch of all species rose from around 100 000 tonnes in 1979 to about 500 000 tonnes in 1989 with the proportion of Nile perch rising from less than 0.1% in 1974 to more than 50% 20 years later. Over the same period the lake lost as many of 50% of its species of haplochromines and became dominated by three species, Nile perch, the introduced Tilapia nilotica, and a native cyprinid (sardine) Rastrineobola argentea. As other changes occurred in the Lake at the same time, such as the advent of the invasive alien water hyacinth (Eichhornia crassipes), it is difficult to attribute changes in diversity of the haplochromines solely to the Nile perch. There are a number of probable causes of the dramatic population increase of the Nile perch: 1. 2. 3. 4. 5.

lack of other wide-ranging predators (competitors); a fast growth rate and reproductive potential; a great range of body size during development which permits exploitation of various habitats in the lake; changes in the lake ecosystem resulting from human activities (e.g., eutrophication); and adaptability of the perch to different sources of food.

It is still not clear though whether or not the population of Nile perch and the wider fish structure in Lake Victoria has reached (or is anywhere near) stability.

difficult to abate, and cyanobacteria blooms have increased and are a major problem worldwide. The extent of water pollution from point sources is well known with an estimated 90% of wastewater in developing countries being discharged directly to rivers and streams without any waste processing treatment, and in some locations both surface- and groundwater have been so polluted that they are unfit even for industrial use (WMO, 1997). The agricultural sector also contributes a large amount, although this is usually from diffuse sources (Verhoeven et al., 2006). It is also well known that pollution from point sources such as mining has had many devastating impacts on inland waters in many parts of the world; for example, the spillage in 1998 of an estimate 5.5 million cubic meters of stored tailings (mine wastes) from the Aznalcollar mine nearly 50 km from the Don˜ana National Park in Spain spread over 46 000 km of downstream habitat with fatal consequences for much of the biota (Bartolome and Vega, 2002). The cost of removing the tailings and contaminated soil reached about 3.8 billion Euros. Meybeck (2003) provided an overview of water pollution problems for inland waters (Table 5). This showed that in industrial countries fecal contamination has been largely eliminated, while new problems, particularly from agriculture runoff, were increasing. In other countries, this was not the case and fecal contamination was still a major problem with urban and industrial pollution sources increasing faster than wastewater treatment. Contamination by pesticides has increased rapidly since the 1970s, with many different substances being involved. Vo¨ro¨smarty et al. (2005) concluded that since the 1990s the water-quality situation in most developing countries and countries in transition was likely to be worse in terms of overall water quality. In Eastern Europe, Central and South populated Americas, China, India, and populated Africa, it was probably worse for metals, pathogens, acidification, and organic matter, while there were slight improvements for the

Table 5 Major water quality issues in inland aquatic ecosystems at the global scale Issue

Rivers Lakes Reservoirs Groundwaters

Pathogens Suspended solids Decomposable organic matter Eutrophication Nitrate Salinization Trace metallic elements Organic micropollutants Acidification

XXXX XXX XXXX

XX NA XX

XX XX XXX

XXX NA XX

XX XX XX XXX

XXX X X XXX

XXXX X XX XXX

NA XXXX XXXX X

XXXX XX

XXX XX

XXX XXX

XXXX X

XXXX, severe or global deterioration observed; XXX, important deterioration; XX, occasional or regional deterioration; X, rare deterioration; NA, not applicable. Information from Meybeck M (2003) Global analysis of river systems: From Earth system controls to Anthropocene syndromes. Philosophical Transactions of the Royal Society London B 358: 1935–1955.

same issues in Western Europe, Japan, Australia, New Zealand, and North America. Nitrate though was generally still increasing everywhere, as it has since the 1950s. In the former Soviet Union there seems to have been an improvement in water quality as a consequence of the decline of industrial activities, whereas in Eastern Europe there have also been some improvements, such as those in the Danube and the Elbe basins. A few rivers, such as the Rhine, have seen a stabilization of nitrate loads after 1995. Arthurton et al. (2007) reported that water-quality degradation from human activities continued to degrade inland aquatic ecosystems and affected the health of many people. Pollutants of primary concern included microbial pathogens and excessive nutrient loads with the latter leading to eutrophication of downstream and coastal waters, and loss of

Managing Aquatic Ecosystems

55

Box 6 Waterbirds and climate change. Based on information reported in Finlayson CM, Gitay H, Bellio MG, van Dam RA, and Taylor I (2006) Climate variability and change and other pressures on wetlands and waterbirds – impacts and adaptation. In: Boere G, Gailbraith C, and Stroud D (eds.) Water Birds around the World, pp. 88–97. Edinburgh: Scottish Natural Heritage. While the general nature of the impacts of climate change on waterbirds can be described there is less certainty when it comes to identifying the extent, intensity, and time frames for such changes. It is difficult to predict with great certainty as the models used for global climate change projections are still very coarse and the ecological relationships between waterbirds and climate and aquatic ecosystems is insufficiently known. The most severe effects and those most likely to occur earliest include: 1. the loss of intertidal areas and increased salinity of coastal freshwater wetlands caused by rising sea levels; 2. a reduction in the extent of wetlands and duration of flooding in arid and semiarid areas from changes in rainfall; and 3. the loss of wetland breeding areas in the Arctic and sub-Arctic areas caused by increasing temperatures, expanding boreal forests and fires.

The extent of loss of intertidal habitats and its effects on coastal waterbirds, many of which also frequent inland aquatic ecosystems, will depend on the ability of coastal environments to migrate inland as sea level rises. The effects of rising temperatures on plant communities will be particularly strong in the Arctic with an expected expansion of the boreal forest into the tundra areas where two-thirds of all goose and 95% of all Calidrid sandpipers breed. The impacts of habitat loss could be offset to some extent by rising temperatures increasing productivity and breeding success; however, these may also be affected by an increase in loss of nests and chicks to predation as predators such as the Red Fox (Vulpes vulpes) expand their range. This example illustrates both the complexity of the changes that may occur as well as the complexity of identifying the many interactions that may occur within an ecosystem or between species. As changes in global circulation patterns will result in changes to rainfall patterns, with some areas experiencing increases and others decreases. The latter in particular may be extremely detrimental to waterbirds in areas that are already dry and subject to drought, such as parts of Australia, Asia, and Africa. As wetlands and waterbirds in these areas are already highly stressed from the impacts of agriculture, reduced water flows, pollution, and increasing salinization, they may be highly vulnerable to changes in the climate. As reduced rainfall will increase the intervals between flooding events and shorten their duration there could be reduced breeding success and recruitment of waterbird species that formerly depended on flooding events of sufficient duration to stimulate breeding and enable fledging. Reduced rainfall and flooding across large areas of arid land will particularly affect bird species that rely on a network of wetlands that are alternately or even episodically wet and fresh or drier and saline. While the exact nature of changes cannot be confirmed there will almost certainly be great regional variation, with some areas experiencing increases in waterbird populations and others, decreases. The fragmentation of rivers and wetlands or the disruption or loss of migration corridors will affect the manner in which waterbirds (and other species) will be able to respond and adapt.

beneficial human uses. Pollution from diffuse land sources such as agriculture and urban runoff was also of concern.

1.03.4.6 Climate Change It is increasingly expected that global climate change will increase the pressure on inland aquatic ecosystems in many locations both directly, especially through increased temperatures, changes in snowmelt and runoff, and in places severely decreasing rainfall, as well as indirectly by interacting with existing pressures and drivers of change (Revenga and Finlayson, 2009; Finlayson et al., 2006; Finlayson and D’Cruz, 2005). Revenga and Finlayson (2009) anticipated that most pronounced impacts from climate change will come from increased temperatures and changes in precipitation, and these will not affect all wetlands in the same way. As a consequence some aquatic ecosystems and their catchments will be drier, while others will experience more rainfall and storms, or even more intense and fewer storms. High-altitude wetlands seem to be particularly vulnerable as the annual and previously predictable glacier-melt decreases; freshwater systems near the coast are susceptible to rise in sea level and salinization. Many combinations of temperature increase and precipitation changes will affect the frequency, duration, and timing of peak floods or base-flows in rivers with subsequent impacts on aquatic species that are sensitive to changes in water flow for migration, breeding, and feeding. Further assessment is needed to ascertain the vulnerability of particular ecosystems and species. The latter has been summarized for waterbirds by

Finlayson et al. (2006) and outlined in Box 6. However, in many instances, the certainty with which we can attribute cause and effect of climate change is undermined by the extent of existing data and knowledge. There is some confidence that many inland aquatic ecosystems are vulnerable to climate change with those at high latitudes and altitudes, such as Arctic and sub-Arctic bog communities, or alpine streams and lakes being highly vulnerable (Gitay et al., 2002; Finlayson et al., 2006; Revenga and Finlayson, 2009), as well as those that are isolated or are low-lying and adjacent to coastal wetlands (Bayliss et al., 1997; Pittock et al., 2001). Danielopol et al. (2003) expected groundwater ecosystems to change as recharge of aquifers is affected by rainfall and runoff. Besides changes in waterbird populations, the warming of inland waters could affect chemical and biological processes, reduce the amount of ice cover and dissolved oxygen in deep waters, alter the mixing regimes, and affect the growth rates, reproduction, and distribution of organisms and species (Gitay et al., 2002). The rise in sea levels will affect freshwater ecosystems in low-lying coastal regions and plant species that are not tolerant to increased salinity or inundation could be eliminated. Changes in the vegetation will affect both resident and migratory animals, especially if these result in a major change in the availability of staging, feeding, or breeding grounds for particular species. The impacts on the distribution of fish species could be profound with cold-water fish being further restricted in their range, and cool- and warm-water fish potentially expanding their range, and even moving poleward. These examples illustrate the statements made at the outset of this discussion that the extent of change in inland aquatic

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ecosystems due to the climate change should not be addressed in isolation of other drivers of change, as many of the adverse effects of the above-mentioned drivers of change will be exacerbated by climate change.

• •

1.03.5 Management Responses The information on management responses is largely paraphrased from that provided by Finlayson and D’Cruz (2005) in their assessment of the condition of inland waters for the Millennium Ecosystem Assessment. As their assessment was drawn from the published literature available at the time, it provides a widespread opinion as well as various options and opportunities for sustainable use and, where necessary, rehabilitation of inland aquatic ecosystems. Particular technical responses for specific wetlands or many drivers of change outlined in the text above are not provided – these are on the whole sufficiently well known, or could be developed through the application of the processes outlined below.

1.03.5.1 Integrated Management Processes The management of inland aquatic ecosystems worldwide has often been based on sectorally based decision-making mechanisms that have not included sufficient consideration of the wider implications or outcomes of specific actions. The information provided in the text above illustrates many adverse outcomes of past sectorally based management decisions. In many instances, these decisions have not adequately considered the trade-offs between the multiple uses and values of inland aquatic ecosystems and have too often resulted in the degradation of these ecosystems. The development of more multi-sectorally based responses and decisions is strongly encouraged as way to reverse the past loss and degradation of inland aquatic ecosystems and the decline in the ecosystem services that they deliver.

1.03.5.2 International Cooperation and Action The past loss and degradation of inland aquatic ecosystems have been recognized through international conventions and treaties. The Ramsar Convention on Wetlands has provided leadership and worked collaboratively with many other organizations, both informally and through formal agreements, to develop multisectoral approaches to stop and reverse the loss and degradation of wetlands. This includes working collaboratively to reduce the rate of loss of biodiversity and restore degraded wetlands. The Mediterranean wetland (MedWet) program is an example of a collaborative initiative that has supported actions to halt and reverse the loss and degradation of wetlands. The declaration behind this initiative was made in February 1991 and contained many recommendations that are still important today. These covered:

• •

identification of priority sites for wetland restoration and rehabilitation, and the development and testing of techniques for their complete rehabilitation; evaluation of existing and proposed policies to determine how they affect wetlands;

• •

increased institutional capacity to conserve and effectively manage wetlands through vigorous education and training programs; integrated management of all activities concerning wetlands, their support systems, and the wider area surrounding them carried out by properly funded and wellstaffed multidisciplinary bodies with active participation of representatives of government, local inhabitants, and the scientific and nongovernmental communities; open consultation and free flow of information when managing wetlands; and adoption and enforcement of national and international legislation for better management.

These recommendations have since been repeated or extended in many forums and with widespread acceptance, although the sentiment behind these have not always been transferred to on-ground actions and outcomes.

1.03.5.3 Restoration and Wise Use of Wetlands The concept of replacing lost wetlands has received increasing support in recent decades and more attention is now directed toward wetland restoration worldwide. However, current rates of restoration are inadequate to offset the rate of wetland loss in many regions – more is required even as efforts are undertaken to stop further loss. In support of these efforts, the Ramsar Convention has provided a suite of guidance for the wise use of wetlands covering national wetland policies; laws and institutions; river basin management; participatory management; wetland communication, education, and public awareness; management planning; international cooperation; wetland inventory, assessment and monitoring; water allocation and management; coastal management; and management of peatlands. One of the key barriers in preventing further loss and degradation of wetlands is the seeming unwillingness of parties to the above-mentioned collaborative initiatives and declarations to undertake effective actions. Sufficient knowledge is generally now available to know what actions are required to stop further loss and degradation, although there seems to be an inadequate adoption and understanding of ecosystem approaches for managing inland aquatic ecosystems, especially when dealing with water allocations. Ongoing dialog about the allocation of water for environmental outcomes in rivers and associated wetlands is still needed – these also need to address the trade-offs that are needed to support equitable outcomes and support many services that inland aquatic ecosystems provide to wider society and the inordinate costs associated with reinstating these once they have been lost.

1.03.5.4 Supporting Local Community Involvement in Management There has been increased interest in the development of mechanisms to support the capacity of local communities to contribute to the management of inland aquatic ecosystems. This can particularly be important where local knowledge and experience can be directly applied to local management issues, but can also support wider strategic planning. Recognition of the beneficial outcomes that can occur when local people are

Managing Aquatic Ecosystems

involved in the management of inland waters and their services has long underpinned efforts by the Ramsar Convention and its partners to encourage best management practices for wetlands. Participatory management and the involvement of local communities in management planning are implicit in the guidance provided by the Convention covering policy and legal instruments, economic and social interactions, and technical tools. The challenge for the Convention and others is to ensure that such instruments and tools are used effectively and as often as possible. This can be done by adopting an adaptive management approach which incorporates active learning mechanisms, the involvement of key stakeholders, and the balancing of vested interests.

1.03.6 Conclusions In drawing the above to a close reference is again made to the Ramsar Convention on Wetlands and the Millennium Ecosystem Assessment. For over 35 years, the Convention has recognized the interdependence of people and their environment. It has promoted the wise use of wetlands as a means of maintaining their ecological character – the ecosystem components and processes that comprise the wetland and underpin the delivery of ecosystem services, such as freshwater and food – and strongly supported the Millennium Ecosystem Assessment with a set of key messages about wetlands and their importance for people. These messages are shortened and paraphrased below as a way of concluding the above review of the condition and management of inland aquatic ecosystems:

•

•

•

•

•

•

Wetlands deliver a wide range of ecosystem services that contribute to human well-being, such as fish and fiber, water supply, water purification, climate regulation, flood regulation, coastal protection, recreational opportunities, and, increasingly, tourism. A priority when making decisions that directly or indirectly influence wetlands is to ensure that information about the full range of benefits and values provided by different wetland ecosystem services is considered. The degradation and loss of wetlands is more rapid than that of other ecosystems. Similarly, the status of both freshwater and coastal wetland species is deteriorating faster than those of other ecosystems. The primary indirect drivers of degradation and loss of inland wetlands have been population growth and increasing economic development. The primary direct drivers of degradation and loss include infrastructure development, land conversion, water withdrawal, eutrophication and pollution, overharvesting and overexploitation, and the introduction of invasive alien species. Cross-sectoral and ecosystem-based approaches to wetland management – such as river (or lake or aquifer) basin-scale management – that consider the trade-offs between different wetland ecosystem services are more likely to ensure sustainable development than many existing sectoral approaches. Major policy decisions in the next decades will have to address trade-offs among current uses of wetland resources

•

•

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and between current and future uses. Particularly important trade-offs involve those between agricultural production and water quality, land use and biodiversity, water use and aquatic biodiversity, and current water use for irrigation and future agricultural production. The adverse effects of climate change will lead to a reduction in the services provided by wetlands. Removing the existing pressures on wetlands and improving their resiliency are the most effective methods of coping with the adverse effects of climate change. The Millennium Ecosystem Assessment conceptual framework for ecosystems and human well-being provides a framework that supports the promotion and delivery of the Ramsar Convention’s wise use concept. This enables the existing guidance provided by the Convention for the wise use of all wetlands to be expressed within the context of human well-being and poverty alleviation.

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Page S, Hosciło A, Woosten H, et al. (2009) Restoration ecology of lowland tropical Peatlands in Southeast Asia: Current knowledge and future research directions. Ecosystems 12: 888--905. Page SE, Siegert F, O’Reiley J, von Boehem H-D, Jaya A, and Limin S (1997) The amount of carbon released from peat and forest fires in Indonesia during 1997. Nature 420: 61–65. Page SE, Siegert F, O’Reiley J, von Boehm H-D, Jaya A, and Limin S (2002) The amount of carbon released from peat and forest fires in Indonesia during 1997. Nature 420: 61--65. Partow H (2001) The Mesopotamian Marshlands: Demise of an Ecosystem, 46pp. Nairobi, Kenya: UNEP. Pittock B, Wratt D, Basher R, et al. (2001) Australia and New Zealand. In: Climate Change 2001. Working Group II of the Intergovernmental Panel on Climate Change: Impacts, Adaptation and Vulnerability, ch. 12. Cambridge: Cambridge University Press. Postel S (1999) Pillar of Sand: Can the Irrigation Miracle Last? New York: WW Norton. Ramsar Convention Secretariat (2006a) The Ramsar Convention Manual: A Guide to the Convention on Wetlands (Ramsar, Iran, 1971), 4th edn. Gland, Switzerland: Ramsar Convention Secretariat. Ramsar Convention Secretariat (2006b) Wise use of wetlands: A conceptual framework for the wise use of wetlands. In: Ramsar Handbooks for the Wise Use of Wetlands, 3rd edn., vol. 1. Gland, Switzerland: Ramsar Convention Secretariat. Ramsar Convention Secretariat (2006c) Inventory, assessment, and monitoring: An integrated framework for wetland inventory, assessment, and monitoring. In: Ramsar Handbooks for the Wise Use of Wetlands, 3rd edn., vol. 11. Gland, Switzerland: Ramsar Convention Secretariat. Ramsar Convention Secretariat (2006d) Designating Ramsar sites: The Strategic Framework and guidelines for the future development of the List of Wetlands of International Importance. In: Ramsar Handbooks for the Wise Use of Wetlands, 3rd edn., vol. 1. Gland, Switzerland: Ramsar Convention Secretariat. Ramsar Strategic Plan 2009–2015. http://www.ramsar.org/pdf/key_strat_plan_2009_e.pdf (accessed August 2010). Rebelo L-M, Finlayson CM, and Nagabhatla N (2009) Remote sensing and GIS for wetland inventory, mapping and change analysis. Journal of Environmental Management 90: 2144--2153. Revenga C, Brunner J, Henninger N, Kassem K, and Payne R (2000) Pilot Analysis of Global Ecosystems: Freshwater Systems. Washington, DC: World Resources Institute. Revenga C, Campbell I, Abell R, de Villiers P, and Bryer M (2005) Prospects for monitoring freshwater ecosystems towards the 2010 targets. Philosophical Transactions of the Royal Society B 360: 397--413. Revenga C and Finlayson CM (2009) Wetlands and climate change State of the Wild. New York: World Conservation Society. Revenga C and Kura Y (2003) Status and Trends of Biodiversity of Inland Water Ecosystems. Technical Series No. 11. Montreal: Secretariat of the Convention on Biological Diversity. Richardson CJ (2008) The Everglades Experiment: Lessons for Ecosystem Restoration. New York, NY: Springer. Richardson CJ, Reiss P, Hussain NA, Alwash AJ, and Pool DJ (2005) The restoration potential of the Mesopotamian Marshes of Iraq. Science 307: 1307--1311. Rieley JO and Page SE (eds.) (1997) Biodiversity and Sustainability of Tropical Peatlands. Cardigan: Samara Publishing. Secretariat of the Convention on Biological Diversity (2006) Global Biodiversity Outlook 2. Montreal: Convention on Biological Diversity. Scanlon BR, Jolly I, Sophocleous M, and Zhang L (2007) Global impacts of conversions from natural to agricultural ecosystems on water resources: Quantity versus quality. Water Resources Research 43: W03437 (doi:10.1029/ 2006WR005486). Scott DA and Jones TA (1995) Classification and inventory of wetlands: A global overview. In: Finlayson CM and van der Valk AG (eds.) Classification and Inventory of the World’s Wetlands, Advances in Vegetation Science 16, pp. 3–16. Dordrecht: Kluwer.

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Sculthorpe CD (1967) The Biology of Aquatic Vascular Plants. London: Edward Arnold. Semeniuk V and Semeniuk CA (1997) A geomorphic approach to global classification for natural wetlands and rationalization of the system used by the Ramsar Convention – a discussion. Wetlands Ecology and Management 5: 145--158. Shepherd G (2004) The Ecosystem Approach: Five Steps to Implementation. Gland, Switzerland and Cambridge: IUCN. The Ramsar ‘Toolkit’, 3rd edn. (2007) The Ramsar Handbooks for the Wise Use of Wetlands. http://www.ramsar.org/cda/ramsar/display/main/ (accessed May 2010). UNEP (2002) Vital Water Graphics – an Overview of the State of the World’s Fresh and Marine Waters. Nairobi: United Nations Environment Programme. UNEP (2006) Challenges to International Waters – Regional Assessments in a Global Perspective. Nairobi: United Nations Environment Programme. UNEP (2007) Global Environment Outlook 4 – Environment for Development. Nairobi: United Nations Environment Programme. UNESCO (2000) Water Related Vision for the Aral Sea Basin for the Year 2025. Paris: United Nations Educational, Scientific, and Cultural Organization. UNESCO-WWAP (2006) Water: A Shared Responsibility. The United Nations World Water Development Report 2. Paris and New York: United Nations Educational, Scientific, and Cultural Organization and Berghahn Books. van Dijk PP, Stuart BL, and Rhodin AGJ (2000) Asian Turtle Trade: Proceedings of a Workshop on Conservation and Trade of Freshwater Turtles and Tortoises in Asia Chelonian Research Monographs, No. 2. Lunenburg: Chelonian Research Foundation in association with WCS, TRAFFIC,WWF, Kadoorie Farm and Botanic Gardens US Fish and Wildlife Service. Verhoeven JTA, Arheimer B, Yin C, and Hefting MM (2006) Regional and global concerns over wetlands and water quality. Trends in Ecology and Evolution 21: 96--103. Vo¨ro¨smarty CJ, Le´veˆque C, and Revenga C (2005) Fresh water. In: Hassan R, Scholes R, and Ash N (eds.) Ecosystems and Human Well-Being: Current State and Trends: Findings of the Condition and Trends Working Group. Washington, DC: Island Press. WCD (2000) Dams and development: A new framework for decision-making. The Report of the World Commission on Dams. London: Earthscan. Whigham DF (2009) Global distribution, diversity and human alterations of wetland resources. In: Maltby E and Barker T (eds.) The Wetlands Handbook. Chichester: Wiley-Blackwell. Whigham DF, Goode RE, and Kvet J (eds.) (1993) Wetland Ecology and Management – Case Studies. Dordrecht: Kluwer. Wiseman R, Taylor D, and Zingstra H (eds.) (2003) Proceedings of the Workshop on Agriculture, Wetlands and Water Resources: 17th Global Biodiversity Forum, 122pp. Valencia, Spain, November 2002. New Delhi, India: National Institute of Ecology and International Scientific Publications. WMO (1997) Comprehensive Assessment of the Freshwater Resources of the World. Stockholm: World Meteorological Organization and Stockholm Environment Institute. WWDR (2003) Water for People, Water for Life. United Nations World Water Assessment Programme. Paris, France: UNESCO/Berghahn Books. Yarrow M, Marı´n VH, Finlayson M, Tironi A, Delgado LE, and Fischer F (2009) The ecology of Egeria densa Planchon (Liliopsida: Alismatales): A wetland ecosystem engineer? Revista Chilena de Historia Natural 82: 299--313.

Relevant Websites http://www.gisp.org Global Invasive Species Programme. http://www.ramsar.org Ramsar: The Ramsar Convention on Wetlands. http://www.ramsar.org Ramsar: The Ramsar Convention on Wetlands; A brief history of the Ramsar Convention.

1.04

Water as an Economic Good: Old and New Concepts and Implications for Analysis and Implementation

J Briscoe, Harvard University, Cambridge, MA, USA & 2011 Elsevier B.V. All rights reserved.

1.04.1 1.04.2 1.04.3 1.04.3.1 1.04.3.2 1.04.3.3 1.04.3.4 1.04.4 References

Introduction Challenge One: Revisiting the Old Issue of the Indirect Effects of Investments in Major Water Projects Challenge Two: Managing Water as a Scarce Resource Issue One: The Radically Different Markets in which Irrigation and Urban Water Operate Issue Two: How Appropriate Pricing Is Understood by Economists and by Users and the Implications for Practice Issue Three: The Crucial Distinction between Financial Costs and Opportunity Costs, and the Implications for Practice Issue Four: The Political Economy of Change Conclusions

1.04.1 Introduction This chapter addresses two conceptual and operational challenges which are of major importance for the management and development of water resources. Challenge one relates to the economic impact of large water infrastructure projects and suggests, first, that conventional economic analytic tools are of little value and, second, that newly emerging tools can be of substantial practical use. Challenge two relates to management and describes both fallacies and emerging approaches for ensuring that the economic productivity of water is maximized.

1.04.2 Challenge One: Revisiting the Old Issue of the Indirect Effects of Investments in Major Water Projects During the era of rapid economic growth, the political leaders of now-rich countries invested heavily in much major water (and other) infrastructure because they believed that these investments would transform the regional economies in which the projects – such as the Tennessee Valley Authority (TVA), Hoover Dam, and Grand Coullee projects in the United States – were located. Decades after these projects were completed (and delivered these large indirect benefits), the US Office of Management and Budget in the 1950s declared that under conditions believed to be prevailing in the United States (full employment and mobile factors of production) these indirect benefits should not be taken into account in investment decisions. As described in detail elsewhere (Briscoe, 2008), this was greeted with incredulity by political leaders who asked ‘‘if we had followed this advice, what infrastructure would ever have been built?’’ (with the implicit answer ‘‘very little’’). Although this issue faded from the sight of politicians, as most of the major infrastructure in the US had already been built, it did not fade from the bible of economists and was incorporated into conventional economic wisdom and standard appraisal practices of institutions such as the World Bank.

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A recent set of detailed analyses (Bhatia et al., 2008) of the Bhakra Dam in northwest India, the Sao Francisco dams in Brazil, and Aswan Dam in Egypt confirm similar findings from earlier studies – namely that such projects have major backward linkages (for inputs into agriculture) and forward linkages (for processing of agricultural products, for instance). In all cases, not only the indirect effects were as large as the direct effects (as had been demonstrated in other analyses of the Muda projects in Malaysia and Grand Coullee in the United States) but also these projects had stimulated precisely the regional development which politicians had hoped for (and which economists now said, ‘‘don’t count’’). Equally important, where the data were available (as in the Bhakra case; Bhatia et al., 2008) it turned out that the biggest proportional beneficiaries were not the landlords but the landless, as a result of the sharp increase in the demand for labor. For decision makers in the real world, the conclusion is that these indirect impacts are large, and that such projects can, indeed, be the basis for regional development. It is true, nevertheless, that there is a serious analytic challenge and an even more serious practical challenge. The analytic challenge is that these studies are all ex post. There is no reliable ex ante method for assessing the indirect impacts. The practical challenge is that there is no established methodology for deciding on what packages of complementary public investments are needed in order to maximize the likelihood that the unquantifiable-but-very-important indirect benefits do, in fact, materialize. On the latter, there is a ray of light, from the original work of Harvard economist Ricardo (e.g., Hausmann and Klinger, 2008) on development patterns as defined by the ‘product space’. This work represents a major intellectual departure from the normative and mechanical work embodied in classic cost–benefit analysis. The approach starts not with principles, but with the collection of data and the use of network approaches to describe revealed patterns of economic development paths in hundreds of economies over time. In the case of major water projects, this would mean the following. First, planners would describe ‘‘where the region is’’ (in our case,

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after the building of the major infrastructure, the emergence of new configurations of energy generation, agriculture, industry, and transport). Second, taking account of the new regional reality, assess the opportunities this starting point affords by examining development paths which have evolved from similar endowments, thus identifying the paths that are most promising (and those which are no more than pie-in-the-sky fantasies). Armed with this X-ray (for the example of Pakistan, see Hausmann and Klinger (2008)), planners of large water infrastructure can then decide in a systematic and informed way on what complementary investments are needed to maximize the likelihood that multipliers will develop. The development of such a methodology is a high priority for developing countries (and there are many), who are in the early stages of investing in major water infrastructure, and offers an escape from what have become ritualistic and uninformative standard cost–benefit procedures.

1.04.3 Challenge Two: Managing Water as a Scarce Resource Many countries face multiple concerns regarding the growing scarcity of water, the associated conflicts among users, and ways of transferring water from low-value to high-value uses. Prominent and well-informed commentators often state that having users pay the full cost of water would solve these problems (recent examples include the CEO of Nestle (Brabeck-Letmathe, 2008) and The Economist (2008)). Experience has shown that the situation is considerably more complex and nuanced, and that it is not enough to just extol the virtues of pricing. This chapter outlines a different approach – one of principled pragmatism. Principled because economic principles such as ensuring that users take financial and resource costs into account when using water are very important; and pragmatism because solutions need to be tailored to specific, widely varying natural, cultural, economic, and political circumstances, in which the art of reform is the art of the possible. The general arguments are illustrated by focusing on two major users – farmers and cities. Here, four issues are addressed. This chapter draws on the World Bank’s Water Resources Sector Strategy (World Bank, 2003): 1. the quite different economic environments that pertain to these two sectors; 2. the crucial distinctions between the perspective of economists and the perspective of users on what constitutes appropriate pricing, and some of the implications of these distinctions for practice; 3. the critical distinction between the financial cost of providing a service and the opportunity cost of the resource itself, and the implications of this distinction; and 4. a review of some good practice developments, and the implications for a country-specific, practical, sequenced approach to dealing with these crucial issues.

1.04.3.1 Issue One: The Radically Different Markets in which Irrigation and Urban Water Operate The first, fundamental distinction is between the markets in which urban water supply and irrigation operate.

In the case of urban water supply, the product can largely be considered as a local, nontradable good. The price charged for water in Helsinki is entirely immaterial to the price charged in Timbuktu. More specifically, if Helsinki chooses to subsidize its water users, that is of no relevance to water users in Timbuktu. In the case of irrigation, where the end products are agricultural goods that trade on a global market, the situation is radically different. If the government of a developed country chooses to subsidize water (and other inputs and outputs) of its farmers, this has an impact on world prices, and thus a direct impact on producers in developing countries. As the magnitude of the agricultural subsidies from OECD countries (OECD, Organization for Economic Cooperation and Development) is huge (about $350 billion/year, to the detriment of consumers in developed countries and producers in developing countries), this has a major impact on the prices of agricultural products in developing countries and on the economic returns from farming. These distortions reinforce the demands of farmers in developing countries with regard to subsidies for water, energy, and other inputs, usually causing further harm to both the economy and the environment. This crucial fact makes the political economy of water pricing reform especially complex (in both theory and practice) for irrigation. Experience suggests that the appropriate approach is to acknowledge the need for subsidies and to document the existing levels. Then it is possible – for example, as has been done in Mexico (Gonzalez, 1997) – for the government and farmers to agree upon a subsidy-neutral transformation from a package of perverse subsidies (of fertilizers, pesticides, and water, for instance) to a package of virtuous subsidies (such as for improving land quality and for more efficient technology).

1.04.3.2 Issue Two: How Appropriate Pricing Is Understood by Economists and by Users and the Implications for Practice Economists have long had a sound theoretical basis for assessing the resource implications of pricing, namely charging users for the marginal cost of producing the next unit of input. This rule is clear and correct, because that is the signal which will cause users to take into account the cost of the next unit of production when they consider using another unit of the resource. Unfortunately, even sound theory does not always translate into rules that can easily be understood and applied in practice. The first reason for this is that ordinary users understand a price as a payment for a service rendered. When the supplier is a monopoly (and prices are set outside of the market), this means that the legitimate price in the eyes of users is that which it costs an efficient producer (usually a public utility) to produce the service. In economic terms, this means that users consider average, not marginal, cost to be legitimate. Two more questions arise from this: What is included in cost and what happens if the service provider is not efficient? Costs that users consider legitimate certainly include, in all cases, the costs of operating and maintaining the existing infrastructure. Moreover, with some explanation and communication, experience (Langford et al. (1999) describe the

Water as an Economic Good: Old and New Concepts and Implications for Analysis and Implementation

Australian case) shows that users see the costs of replacement as legitimate costs. However, even under the most advantageous of settings, users vigorously resist the notion that they should pay for sunk costs which, in their eyes, have already been paid for by taxes or other assessments. The issue of the efficiency and accountability of the service provider is critical. ‘‘Why should I pay the costs of the Water Department when it is overstaffed, corrupt, and does not maintain our systems?’’ is a frequent and legitimate complaint from consumers and farmers. An illustration of the lower bound of these inefficiencies comes from the state of Victoria in Australia. Before reform, irrigation services were provided by a government department with well-trained and wellperforming staff, and there was little corruption. When reform took place, and farmers had to pay the full costs of operation and maintenance, increased scrutiny of the supply agency led to a 40% reduction in these costs. In most developing countries, the inefficiency is much greater and the users’ resistance to paying for these services is correspondingly higher. Exhortations to increase cost recovery without addressing these fundamental accountability questions are a major part of the reason why cost recovery has been so poor in many countries. A review by the World Bank’s Operations Evaluation Department (2003) shows that, despite the fact that the World Bank has been by far the most constant and insistent advocate of cost recovery for decades, ‘‘there is no evidence of better cost recovery or of covenant compliance either.’’ The bottom line, then, is that in most urban and irrigation systems cost recovery is critical for the supply of good services. The road to cost recovery does not lie in conditionalities imposed by aid agencies, however, but in realigning the institutional arrangements so that suppliers are accountable to users, and so that charges become a principal tool used for ensuring the mutual obligations of suppliers and users.

1.04.3.3 Issue Three: The Crucial Distinction between Financial Costs and Opportunity Costs, and the Implications for Practice User payments for the financial costs of services rendered is a fundamental requirement for any financially sustainable water supply system – this is very important. However, the claims for pricing typically go beyond that of maintaining and operating infrastructure, and suggest that if ‘‘the prices are right, allocation will be optimal.’’ Proceeding from the viewpoint of users (as one must when considering political economy of reform rather than theoretical elegance), it is vital to distinguish between two radically different types of costs. First, there are the costs that any user can understand, namely the financial costs associated with pumps, treatment plants, and pipes. Second is the far more subtle concept of the opportunity cost of the resource itself. There have been many proposals for doing sophisticated calculations of this opportunity cost, and charging users for this ‘‘to ensure appropriate resource allocation.’’ This has not worked in practice for three fundamental reasons: first, because it is impossible to explain to the general public (let alone to angry farmers) why they should pay for something that does not cost anything to produce; second, because opportunity costs vary widely by place and time and could not be

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accurately calculated by even the most sophisticated of regulatory agencies; and, third, because those who have implicit or explicit rights to use of the resource (correctly) argue that they have already paid for the (implicit or explicit) rights and argue (appropriately) such proposals to be the confiscation of property. An added, and highly relevant, factor is that the ratio between financial and opportunity costs is often radically different for different sectors (Briscoe, 1996). Although everything in water (like politics) is local, there are two broad patterns. It costs a lot (per unit of water) to operate the dams, water and wastewater treatment plants, and pumps and pipes that provide households with the modest amount of water they use (and the sewage that is removed). Alongside these large financial costs, the opportunity cost of the resource itself (as measured by the value of the raw water in its next best use, often irrigation) is typically quite low. For municipal and industrial water, therefore, financial costs generally dominate opportunity costs. For irrigation, the situation is almost exactly the opposite. It costs relatively little (per unit of water) to build, operate, and maintain the usual gravity systems that provide very large quantities of water. However, the opportunity cost of the water (for cities and, increasingly, for high-value agricultural uses) is, in situations of scarcity, often much higher (typically at least an order of magnitude higher) than the financial cost of supplying the water. These numbers (remembering, of course, that every place is different) have profound implications. They mean that, from the point of view of ensuring that users take into account the cost of the resources they are using, the emphasis must be on financial costs for municipal supplies, and on opportunity costs for irrigation. (It is worth emphasizing that this does not mean that cost recovery does not matter for irrigation. Cost recovery for irrigation remains very important for infrastructure sustainability, but not for efficiency in the allocation or use of water.) The great challenge for irrigation, in light of these theoretical and practical realities, is how to have farmers take account of the opportunity cost of water. In most parts of the world where water is scarce, informal water markets have arisen, in which those who have (implicit) rights sell water to those who need it. In some cases, the practice has existed for hundreds of years and has been formalized (as in the Water Court of Valencia, Spain, which has managed transfers among users for a 1000 years). In many other cases (such as western India; Shah, 1993), these markets are extensive, sophisticated, and illegal. Throughout the arid western United States, water rights have long been legal property and, under different rules in different states, allowed for approved transfers between willing buyers and willing sellers. As other parts of the world have experienced scarcity, a number of countries facing water stress have turned toward formal, legal, managed water markets. This took place in recent decades in Chile, Australia, and Mexico. The Australian case shows the benefits – the adverse impact of water reductions on the regional economy – are reduced by two-thirds when there is both intra- and interstate water trading (Productivity Commission, 2004).

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Water as an Economic Good: Old and New Concepts and Implications for Analysis and Implementation

From the perspective of the present discussion on how to ensure that users take account of opportunity costs, these market-based arrangements have a unique virtue. Once users have clear, transferable property rights, then they automatically consider whether they wish to forego a particular use of water in exchange for compensation from another user who may place a higher value on the water. Reallocating water then becomes a matter of voluntary and mutually beneficial agreements between willing buyers and willing sellers, and not a matter of confiscation or an endless search for new sources of supply. This is not to suggest that the establishment of water markets is simple or a panacea. The operation of such systems is demanding in terms of rules for establishing initial rights (including those for the environment and informal customary rights); the plumbing required to measure and move water; the regulatory institutions that are essential to protect the rights of other water users and the environment and to ensure that the public interest is represented; and the information and management systems. Many consider these prerequisites so onerous that they cannot be made to work in most developing countries. In addition, many point to early problems that all countries have faced in making such changes. Without in any way minimizing these challenges, three observations are germane. First, the prerequisites are really prerequisites for any form of well-managed allocation system and the absence of such prerequisites is a problem for all allocation systems, including the administrative allocation systems practiced in most countries. (As with everything in water management, the choice is not between the first and the second best, but between imperfect and even more imperfect.) Second, one of the many virtues of a market-based system is that, once started, there is a strong demand for better measurement, transparency, regulation and information. Third, all such established systems are working, often after initial adjustments, reasonably well. In none of the countries that have adopted such systems is there any thought to reverting to the previous allocation procedures.

1.04.3.4 Issue Four: The Political Economy of Change The implications for practitioners are clear. First, from the point of view of financial cost recovery, the key is an institutional framework whereby service providers are accountable and efficient. When this materializes, and when users see that their payments are being used to improve the quantity and quality of services, they can and will pay. Here (as discussed earlier), watchwords are competition, regulation, transparency, benchmarking, and accountability. In the urban water supply and energy sectors, these ideas are now accepted in most parts of the world. In the irrigation sector, there is a gradual, albeit still far too slow, acceptance of these principles. Building on the historic experiences in countries such as Spain and the United States, a number of countries (including Australia, Chile, Mexico, and, more recently, the provinces of Punjab in Pakistan (Government of Punjab, Pakistan, 2008) and Maharashtra in India (Government of Maharashtra, 2003)) have moved toward systems which (1) charge irrigators for the cost incurred in providing services and (2) have clarified and made transparent water entitlements which will,

slowly and inexorably, lead to trading and the revelation of opportunity costs. In all settings, a critical element of this approach is to develop innovative mechanisms for breaking out of the typical low-level equilibrium, in which services are poor, users will not pay, service quality declines, etc. In one good example of such innovation, the World Bank helped the government of Guinea Conakry break the circle by guaranteeing a new, accountable operator a declining proportion of reasonable costs over a 5-year period (World Bank, 1993). In the first year, then, the operator had sufficient revenues (mostly from the International Development Association (IDA) credit, but some from users) to improve the operation of the system. As the level of service improved, users were informed that they would be charged for the new, improved service and that, eventually, they would pay the full costs of the service. The art of reform is less one of articulating a vision than of tracing a path for making improvements, for applying generic principles in a way that takes account of the very widely varying historical, cultural, natural, social, and economic conditions which govern water management (Briscoe, 1997). An analysis of experiences of successful reforms suggests that this means, inter alia: ‘‘picking the low-hanging fruit first,’’ for instance, by starting with temporary trading in well-defined systems where good infrastructure is in place; ‘‘not making the best the enemy of the good,’’ by having a well-defined, sequenced, prioritized, and patient approach for moving toward improvement, not seeking to attain perfection in one fell swoop; and ‘‘keeping one’s eyes peeled,’’ by understanding that it is broader reforms outside of the water sector (often relating to overall economic liberalization and fiscal and political reform) which will provide the preconditions for making the critical first steps. Recent reviews of water reforms in Pakistan and India (Briscoe and Qamar, 2007; Briscoe and Malik, 2006) describe, in considerable detail, what the application of these principles might be in practice.

1.04.4 Conclusions There is growing understanding that there are broad benefits – for the economy, users, and for the environment – if water is developed and managed as an economic good and a growing search for a new set of analytic and operational tools. In recent years, there has been a subtle but important change in discussion of economic policy. The landmark Growth Commission (Spence et al., 2008), written by several Nobel prize laureates and many eminent development practitioners, draws lessons from the history of successful growth experiences. The Commission discarded the rigid prescriptions so often advocated, and noted that there were a wide variety of different, successful, experiences. What they did conclude was that there were some common elements – for example, a disciplined examination of, and adherence to, comparative advantage – and then application of economic principles in a sequenced, nuanced manner appropriate to particular cultural and economic circumstances. Application of this less rigid approach has major implications for water.

Water as an Economic Good: Old and New Concepts and Implications for Analysis and Implementation

In terms of the development of infrastructure, it means getting away from what have become uninformative, formulaic analyses of internal rates of return, to an approach which uses new tools to identify critical supplementary investments needed to maximize the multiplier effects of major investments. In terms of management, it means moving away from the tired phrase of ‘get the prices right’ (and everything will be okay) that has been repeated for years, with little impact on the ground. It means paying much more attention to incentives and to opportunity costs, and to creating an enabling environment in which users will make much better use of limited water, or transfer the right to use that water to others who can use it more productively. Finally, it also means giving greater attention to the political economy of change. This chapter advocates (as does the 2003 World Bank Water Resources Strategy) a path of principled pragmatism – in which the principles of sound economic management are well defined and respected, but in which they are applied in a pragmatic and sequenced way which takes account of local circumstances and political economy, and in which the focus is on moving in the right direction, and on the art of the possible.

References Bhatia R, Cestti R, Scatasta M, and Malik RPS (eds.) (2008) The Indirect Economic Impact of Dams. New Delhi: The Academic Foundation. Brabeck-Letmathe P (2008) Global drying. Wall Street Journal, Asia 13 June 2008: A13. Briscoe J (1996) Water as an economic good: The idea and what it means in practice. In: Proceedings of the World Congress of the International Commission on Irrigation and Drainage. Cairo, Egypt, September 1996.

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Briscoe J (1997) Managing water as an economic good: Rules for reformers. Water Supply 15(4): 153--172. Briscoe J (2008) How theory, practice, politics and time affects views on the indirect economic impact of water infrastructure. In: Bhatia R, Cestti R, Scatasta M, and Malik RPS (eds.) The Indirect Economist Impact of Dams. New Delhi: The Academic Foundation. Briscoe J and Malik RPS (2006) India’s Water Economy: Bracing for a Turbulent Future. New Delhi: Oxford University Press. Briscoe J and Qamar U (2007) Pakistan’s Water Economy: Running Dry. Oxford: Oxford University Press. Gonzalez F (1997). Water Reforms in Mexico. Water Week, World Bank. Government of Maharashtra (2003) Maharashtra Water Regulatory Authority Bill XIX. Mumbai. Government of Punjab, Pakistan (2008) Entitlements. http://irrigation.punjab.gov.pk/ Entitlement.aspx (accessed July 2010). Hausmann R and Klinger B (2008) Structural Transformation in Pakistan. Center for International Development, Harvard University. Langford KJ, Foster CL, and Malcolm DM (1999) Towards a Financially Sustainable Irrigation System, World Bank Technical Paper 413. Washington, DC: World Bank. Operations Evaluation Department (2003) Bridging Troubled Waters: Assessing the Water Resources Strategy Since 1993. Washington, DC: World Bank. Productivity Commission (2004) Modelling Water Trade in the Southern Murray Darling Basin, Staff Working Paper, Canberra. Shah T (1993) Groundwater Markets and Irrigation Development: Political Economy and Practical Policy. Bombay: Oxford University Press. Spence M, et al. (2008) The Report of the Growth Commission: Strategies for Sustained Growth and Inclusive Development. Washington, DC: World Bank. The Economist (2008) Running dry. The Economist 18 September 2008. United States Congress (1955) Discussion of Budget Bureau Circular A-47, Hearings before the Committee on Interior and Insular Affairs, Serial no. 5. World Bank (1993) Development and Environment: The World Development Report. Washington, DC: World Bank World Bank (2003) The World Bank’s Water Sector Strategy. Washington, DC: World Bank.

1.05 Providing Clean Water: Evidence from Randomized Evaluations A Ahuja, Harvard University, Cambridge, MA, USA M Kremer, Harvard University, Cambridge, MA, USA AP Zwane, Bill and Melinda Gates Foundation, Seattle, WA, USA & 2011 Elsevier B.V. All rights reserved.

1.05.1 1.05.2 1.05.2.1 1.05.2.2 1.05.3 1.05.3.1 1.05.3.2 1.05.4 1.05.4.1 1.05.4.2 1.05.4.3 1.05.4.4 1.05.5 1.05.6 1.05.6.1 1.05.6.2 1.05.6.3 1.05.7 References

Introduction Water Quantity Health Impacts Maintenance Solutions Water Quality Health Impacts Valuation Nonprice Determinants of Clean Water Adoption Information on Water Contamination Levels Gain versus Loss Framing and Other Behavioral Marketing Communal versus Individual Persuasion Personal Contact Potentially Scalable Approaches to Improving Water Quality Methods and Theory: Contributions of Randomized Evaluations of Domestic Water Survey Effects Valuation: Revealed Preference versus Contingent Valuation Combining Randomized Evaluations with Structural Modeling Conclusion

1.05.1 Introduction Some 1.6 million children die each year from diarrhea and other gastrointestinal diseases for which contaminated drinking water is a leading cause (Wardlaw et al., 2010). This chapter critically reviews experimental work on the provision of water and improved water quality for domestic use in developing countries, discussing both policy implications and methodological lessons. Earlier work has been reviewed in research using nonrandomized approaches (Zwane and Kremer, 2007). Holla and Kremer (2008) provide a summary of the literature on randomized evaluations related to pricing and access in health and education. Cardenas (2009), Pattanayak and Pfaff (2009), and Timmins and Schlenker (2009) provide reviews on related issues. Recent calls for investment in experiments in environmental economics include Greenstone and Gayer (2009) and Bennear and Coglianese (2005). Local public good investments for services such as water for domestic use are arguably typically best prioritized by local policymakers who know the preferences of the communities they serve. However, the sole quantitative environmental target in the United Nations Millennium Development Goals is the call to ‘‘reduce by half the proportion of people without sustainable access to safe drinking water.’’ In practice, efforts to meet this goal have translated into increased donor and national government funding for building local public goods such as wells and standpipes. By increasing the number of water points, this reduces the time to collect water and makes the task more convenient.

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Health externalities that could cross jurisdictional boundaries could be an important part of the case for national or supra-national investments targeted specifically at the water sector. Diarrhea is an infectious disease. Distributional concerns provide another potential rationale for national or international policymakers to target aid to the water sector in particular. Outsiders may place more value on the consumption of child survival goods relative to consumption of other goods than the local household or other local decision maker does. One of the leading debates in the literature has been on the relative health impact of increases in water quantity versus improved water quality. Simply providing more convenient access to water, even without improving water quality, could potentially stimulate greater handwashing, which has been shown to be very important for health, and more washing of clothes and dishes. At this point, however, the limited evidence available from randomized studies does not demonstrate that increasing access to water without changing its quality improves health. In contrast, there is now abundant evidence that improved water quality reduces self-reported diarrhea. This evidence is an example of how randomized evaluations, which often yield quite different estimates of impact than nonexperimental analyses, can clarify questions that are difficult to resolve using analytic techniques that have difficulty separating causal effects of programs from selection bias. As seen below, there is some reason for external support for water quantity beyond what might be chosen by local decision makers, but much less than for water quality. Though there is

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limited evidence from randomized trials for health (as opposed to a time use or convenience) benefits of water quantity, increasing water quantity disproportionately benefits women and thus there may be a distributional case for public policy to support increased water quantity. In contrast, there is strong evidence that improving water quality improves health by reducing infectious disease and thus builds a case for subsidizing water-quality treatment based on reducing externalities. Improving water quality disproportionately also helps an even more severely under-represented group: young children. As discussed below, the evidence of low valuation for improvements in water quality, in contrast to water quantity, is consistent with the hypothesis that households put very little weight on the health costs of dirty water in the case of young children. External funders or national policymakers may put more weight on child health. A challenge with prioritizing water-quantity investments for donor funding is that maintaining the associated hardware has traditionally been difficult. When infrastructure falls into disrepair, the stream of benefits associated with the investment may be lost. Randomized impact evaluations suggest that external contracting can perform better than community-based voluntary arrangements at least in some circumstances. Evidence from India suggests that women are more likely to invest in water infrastructure but evidence from Kenya suggests little effect of efforts to encourage selection of female user committee chairs on quality of water infrastructure maintenance. Water-quality interventions face a different challenge than hardware maintenance if their biomedical benefits are to be sustained over time. In contexts where treatment does not occur at a centralized treatment plant, as in most developing countries, individuals influence the level of diffusion and adoption of treatment technologies. Randomized impact evaluations have provided evidence on the determinants of uptake in such cases, including a steep demand curve for treatment products such as chlorine or water filters. In addition to shedding light on policy debates such as the investment decision regarding quality versus quantity, experiments can help researchers come up with new solutions to hurdles such as this technology adoption decision. Randomized evaluations have demonstrated that the demand curve can be shifted outward by providing information and making treatment easy and convenient, as well as local promotion of ongoing use. Combining evidence of low valuation with this other information about influencers of adoption has allowed for new approaches to the service delivery problem to be developed. In particular, providing dilute chlorine solution free at the point of water collection, together with a local promoter, can increase takeup of water treatment from less than 10% to more than 60%. Methodologically, randomized evaluations have provided evidence that the process of collecting data through surveys can itself affect behavior and that revealed preference estimates of willingness to pay for environmental interventions in developing countries are far smaller than stated preference estimates. Recent work also marries randomized evaluations with structural modeling to provide guidance on the potential impact of alternative policies and social norms. The remainder of this chapter is structured as follows: Section 1.05.2 summarizes evidence from randomized

evaluation on the impact of infrastructure investments to increase water-quantity improvements in developing countries and on the maintenance of these investments. Section 1.05.3 argues there is considerable evidence that water-quality improvements yield health benefits, but that many households are willing to pay very little for cleaner water. Section 1.05.4 assesses alternative means of shifting the demand curve for water quality, examining information provision, communal versus individual persuasion, and local promoters. Section 1.05.5 discusses cost-effective and potentially scalable approaches to water quality drawing on the lesson of Sections 1.05.2, 1.05.3, and 1.05.4. Section 1.05.6 reviews methodological contributions from randomized evaluations of domestic water interventions. Section 1.05.7 concludes this chapter.

1.05.2 Water Quantity 1.05.2.1 Health Impacts Identifying the aspects (quantity vs. quality) of improved water supply is important for policy because different interventions affect quality and quantity asymmetrically. For example, adding chlorine to water affects quality but not quantity. Providing household connections to municipal water supplies to households that currently use standpipes is likely to have a bigger effect on the convenience of obtaining water and thus on the quantity of water consumed than on water quality. There has been considerable debate over whether increasing the quality of water or increasing the quantity of water has a greater impact on disease. Much of the most convincing nonexperimental evidence on the health impact of water and sanitation makes it difficult to separate the impact of quantity and quality (Cutler and Miller, 2005; Watson, 2006; Galiani et al., 2005; Gamper-Rabindran et al., 2010) because the interventions that are studied both reduced the cost of collection and improved quality, making it unclear which route of disease transmission mattered the most in practice. In the 1980s and 1990s, nonrandomized studies were frequently cited as evidence that water-quantity interventions were more important for health impacts than water-quality interventions (Esrey, 1996; Esrey et al., 1991). Some argued that these results could be explained because, when water supplies are rationed, increased availability of and convenience of water facilitate more frequent washing of hands, dishes, bodies, and clothes, thus reducing disease transmission (Esrey, 1996; Esrey et al., 1991; Curtis et al., 2000). However, the question remained unsettled because it was difficult to assess causality in the absence of randomized evaluations or other convincing identification. In the past 10 years, a new body of evaluations has been developed to address this question. We discuss in Section 1.05.3 the numerous randomized evaluations that have shown impacts of improved water quality on health while also confirming the importance of hand washing in reducing disease transmission. There remains limited evidence on the question of water quantity, with one recent relevant randomized evaluation. Consistent with the early claims made about the relative importance of water quantity, health benefits from hand

Providing Clean Water: Evidence from Randomized Evaluations

washing have been shown in several settings. Luby et al. (2004) reported the results of a cluster-randomized trial in a large sample of households in Karachi, Pakistan, of a hand washing promotion campaign aimed at mothers. Infants and malnourished children under age 5 living in treatment households had 39% fewer days of diarrhea compared with the control group after 1 year of intervention and observation. Two other older randomized controlled trials of hand-washing interventions with more than two communities in their samples (Khan, 1982; Han and Hlaing, 1989) each had a relatively large sample size randomly divided into treatment and control groups and measured compliance by observing or weighing provided bars of soap as well as by tracking diarrhea cases. The studies report large effects of hand washing and soap provision programs on the incidence of diarrhea. Khan (1982) reported that the provision of either soap and water-storage containers or soap alone, along with initial instructions to increase the frequency of hand washing, reduced shigella reinfection by 67% in Bangladesh. Han and Hlaing (1989) reported a 40% reduction in diarrhea incidence among children under age 2 (though there was no reduction in incidence for older children) following hand washing education and the provision of soap to a random sample of mothers in Rangoon (Yangon). Although impacts may be heterogeneous across settings, and caution is warranted in drawing general conclusions, the one available randomized evaluation found that increasing the quantity of water while maintaining unchanged quality did not lead to significant health improvements. DeVoto et al. (2009) examined provision of piped connections to homes in urban Morocco previously served by standpipes. This increased the quantity of water used by the household, but did not improve water quality, since the alternative was chlorinated water from communal taps, which was of similar quality to the water received at home. As part of a planned piped water service extension in Tangier, Morocco, the authors randomly selected half the households eligible for a first connection to receive information about and an offer of credit toward a new connection, and administrative assistance in applying for credit. Takeup was 69% (as compared to 10% in the control group). The authors compare outcomes of those who received this treatment to those for households in the control group. They find that piped water provision in this urban Moroccan context had few health benefits. There is no evidence for an impact of treatment on a subjective ranking of health of the family or on diarrhea in children under age 6 (though baseline rates were relatively low, with the average child in the control group experiencing 0.27 days of diarrhea in the past week). Households in the treatment group report increasing their frequency of baths and showers: the number of times respondents in the treatment group washed themselves (baths, showers) during the last 7 days is 25% higher than in the control group. However, hygiene practices that require less water, such as hand washing, were not affected, according to self-reports. We would not conclude that increased water quantity never yields health benefits. The benefits of increased water quantity may be context specific and require further research to fully understand. In particular, understanding when and how

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increased access to water leads to more hand washing is a research priority. Having a piped water connection had substantial private benefits, despite the lack of impacts on self-reported diarrhea, consistent with the evidence of private valuation. In particular, it saved time, which was used for leisure and social activities. Evidence of substantial willingness to pay for water quantity has been noted by other authors as well in observational studies. In Morocco, the intervention also improved measures of social integration and overall welfare for households. Consistent with this finding, households are willing to pay a substantial amount of money to gain access to a private tap at home: 1 year into the program, not only had the encouragement design resulted in high rates of takeup in the treatment group, but also, for these households, their average monthly water bill more than doubled, from 73 to 192 Moroccan dirhams (MAD), or US$9 to $24 a month (the previous cost came from households, who took water from their neighbors). The importance of a household visit in inducing adoption is something we discuss further in Section 1.05.4.4; additional evidence on personal contact has been generated in other settings as well. There is evidence from India that women particularly value water investments and that women’s involvement in investment decisions could result in improved water supply in situations where local governments set priorities among local public good investments. Women’s valuation of these goods can be part of a distributional argument in favor of additional external support for these investments as well. Chattopadhyay and Duflo (2004) found that a randomized policy change in India that increased the role of women in policy decision making led to more investment in water infrastructure. A 1993 constitutional amendment called for one-third of village council leader positions to be reserved for women. Rules ensured random assignment of the leadership reservations. Chattopadhyay and Duflo showed that village councils headed by women were significantly more likely to invest in public infrastructure for drinking water. These investments were borehole wells and other storage infrastructure that likely improve quality as well as the convenience of water collection. We takeup in the next section the question of whether women’s involvement in local decision making can also make these kinds of investments easier to maintain and keep up.

1.05.2.2 Maintenance Solutions In general, water-quantity investments, whether bundled with water-quality improvements or not, often require significant infrastructure investments. Water quality can be improved with virtually no investment in infrastructure (e.g., by leaving water in the sun or the addition of chlorine), though of course other water-quality interventions can also require hardware (as the first example in Section 1.05.3 illustrates). Along with infrastructure investments comes the challenge of maintenance, which has historically been a major problem in developing countries. The rural water sector in particular has a poor track record of maintaining infrastructure investments. For instance, a quarter of India’s water infrastructure is believed to be in need of repair (Ray, 2004). World Development Report 2004

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(World Bank, 2003) estimates that more than a third of rural water infrastructure in South Asia is not functional. Miguel and Gugerty (2005) reported that nearly 50% of borehole wells dug in a large project in western Kenya in the 1980s, and subsequently maintained using a community-based maintenance model, had fallen into disrepair by 2000. Difficulties with maintaining water infrastructure, particularly in rural areas, reduce the cost effectiveness of these interventions relative to other measures that prevent diarrhea. There are two solutions frequently mentioned as potential elements of a solution to this infrastructure challenge: empowering women to manage water resources and including communities in participatory management schemes. As discussed above, women value water investments when capital expenditure decisions are made. Randomized evidence on these solutions suggests that neither may be as effective as contracting out services via maintenance contracts, however, as we review below. Kremer et al. (2008) provided evidence from a randomized field experiment in Kenya that, at least in that context-enhanced women’s involvement in infrastructure management, did not lead to better maintenance of water supplies. This evaluation studies the impact of female affirmative action policies on actual management outcome measures relevant for protected springs (e.g., time since storm drains or drainage trenches were cleaned). When protected springs were provided to 100 communities in rural Kenya, all communities formed water-user committees. In addition, onehalf of the communities received messages encouraging women to take leadership roles in their water-user committees. This encouragement intervention did result in more women representation in the treatment communities. Communities that received the female participation intervention were twice as likely to have women in the role of water committee chair. However, this did not lead to differences in the effectiveness of the user committees’ spring management as measured by the maintenance outcome variables. Thus, the authors conclude that advocacy for female participation can increase women’s involvement without any impact (either positive or negative) on project outcomes. This has a positive interpretation: empowerment goals can be met without attended costs to project outcomes, as well as a more negative one: including women in management cannot alone solve the water infrastructure maintenance challenge, even if these investments are priorities for women. In addition to increasing women’s participation and decision-making power, another standard model for maintaining donor-funded infrastructure projects, such as water schemes, is to establish user groups responsible for maintenance and management. This approach grew out of the widespread perception that centralized government maintenance was unsuccessful. Giving communities direct control or ownership over key project decisions was intended to improve the quality of public services and increase financial sustainability. There is little convincing empirical evidence, however, that local user-committee management of local public goods such as improved drinking water sources results in better quality service than other models ongoing centralized funding from public budgets. Collective action problems may be difficult to overcome, and voluntary committees tasked with collecting user fees may be difficult to sustain or empower. In a recent

comprehensive review of community-based development projects, Mansuri and Rao (2004) noted that existing research examining successful community-based projects does not compare these projects with centralized mechanisms for service delivery or infrastructure maintenance (e.g., city or state financed). This makes it difficult to determine whether alternative project designs would have had different results. The limited empirical evidence suggests that the impact of the community-based development approach on infrastructure maintenance is mixed at best. In addition to randomly assigning the gender empowerment encouragement intervention, in the same study as described above, the nongovernmental organization (NGO) randomly assigned communities to contracted maintenance and community-based management schemes. Kremer et al. (2008) compared payments to private contractors for spring maintenance and ongoing grants to user committees, with the outcomes of a control group, in which user committees received no grants. The traditional model, user committees without grants, performed worse than either alternative across a range of maintenance outcomes. Providing grants to user committees improves a measure of overall water source maintenance quality by around 30% of one standard deviation on average, while paying contractors to maintain water source leads to an average improvement in measured maintenance quality by around 50% of one standard deviation. This difference is significant at the 10% level. This evidence from spring protection maintenance, a relatively simple technology that seems favorable to communitybased management, suggests that contracting for private maintenance service may be a promising alternative to committee-based management schemes. Nonexperimental evidence from Argentina (Galiani et al., 2005) also suggests that contracted private provision of service can expand coverage and improve health outcomes at least in certain settings in middle-income countries. Certainly, further research is needed that transparently compares the counterfactual of subsidized public service provision and community-based management schemes. In summary, the health benefits of water-quantity interventions require further investigation. Increasing availability of water, even leaving quality unchanged, brings major nonhealth benefits; yet insofar as these seem unlikely to create externalities beyond the household, let alone cross-jurisdictional externalities, local governments may be the proper institution for allocating budgets between water and other public goods. There may be a distributional case for national or supra-national water investments, as these are valued by women, however. Whatever benefits water-quantity interventions do provide can quickly be lost if infrastructure falls into disrepair or is broken. Contracting models seem to hold promise for maintaining water infrastructure.

1.05.3 Water Quality 1.05.3.1 Health Impacts A body of randomized evaluations examines interventions that affect water quality without affecting water quantity. These yield strong evidence of reductions in reported diarrhea.

Providing Clean Water: Evidence from Randomized Evaluations

One study examines source water-quality improvements. Kremer et al. (2009a), in the first randomized evaluation of the provision of improved communal water infrastructure, estimated that protecting springs reduced fecal contamination as measured by the presence of E. coli bacteria by two-thirds in water at the source, but only by 25% for water stored at home. This is likely in part due to recontamination in transport and storage within the household, as well as to the use of alternative sources. Despite the incomplete pass through of the water-quality improvement, this led to a reduction in selfreported child diarrhea of about 25%. Other epidemiological evidence on water quality, manipulated via filtration or treatment rather than infrastructure, also suggests that there may be large health gains from investments in quality. The bulk of the evidence suggests that, with takeup rates on the order of 70% (achieved via frequent visits and reminders to subjects) household water treatment reduces child diarrhea by 20–40%. (One caveat is that the outcome measure in these studies is typically mothers’ reports of child diarrhea. Studies with objective outcomes, infrequently measured, would be desirable (Schmidt and Cairncross, 2009). Nonetheless, we believe that the weight of the evidence is strong enough to believe that reductions in diarrhea are real. To the extent that reporting bias lowers estimates of diarrhea in both the treatment and comparison groups, it may in fact make it harder to pick up reductions in diarrhea. The extent of reporting bias in treatment groups would have to be very large to explain the reported reductions in diarrhea associated with cleaner water. If the reductions in diarrhea are even a fraction as large as those estimated, water treatment would still be very cost-effective.) Comprehensive reviews of this literature are provided by Waddington and Snilstveit (2009), Fewtrell et al. (2005), Arnold and Colford (2007), and Clasen et al. (2006). Because water treatment can be extremely cheap, even a 20–40% reduction in diarrhea makes water treatment very low cost per disability adjusted life years saved. To get a sense of how cheap it is to treat water, note that 1.42 gallon generic bottle of bleach with approximately 6% sodium hypochlorite concentration sold in Walmart for $2.54 as of December 2009 has enough chlorine to treat 163 400 l of water. This corresponds to a price of $0.00002 per liter of water treated. Even making generous allowances for the fact that chlorine used for water treatment should be sold at lower concentrations and has to be transported, etc., if mortality reductions are proportional to reported morbidity reductions, the cost per DALY is similar to that of childhood vaccination, at under $40 per DALY.

1.05.3.2 Valuation Despite the evidence of health benefits associated with water quality, a number of papers suggest very little willingness to pay for this class of interventions. Moreover, there is little evidence that households with young children place substantial additional value on clean water, suggesting low valuation of child health. Kremer et al. (Spring cleaning: rural water impacts, valuation, and institutions, unpublished manuscript) exploited exogenous changes in the trade-off that households face when choosing between multiple water sources, some of which are

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close but contaminated and others of which are far but clean. This variation in the distance/water quality trade-off is generated by the spring protection intervention discussed above that was randomly phased in to almost 200 communities in rural Kenya. (Spring protection reduces contamination by sealing off the eye of the spring so that it is no longer vulnerable to surface-water runoff.) The authors compare how many trips households make to protected springs and other sources, controlling for differences in the time it takes to walk to each source. The estimated mean valuation for spring protection is equivalent to 32.4 workdays. Based on household reports of trade-offs between walking time and money, this corresponds to approximately US$2.96 per household per year. Under additional assumptions, this translates into a willingness to pay $23.68 per DALY saved, which is well below the benchmark of $100–150 often assumed to be appropriate for health investments in developing countries. Kremer et al. (2009a) used randomly assigned discounts to investigate willingness to pay for dilute chlorine. They described behavior consistent with a steep demand curve for water treatment and found no evidence of higher valuation among households with vulnerable young children. In a set of impact evaluations that tested both price and nonprice interventions to increase takeup of chlorine, households were randomly assigned either to a comparison group or to treatment arms in which they received a free supply of individually packaged chlorine or coupons for half-priced chlorine that could be redeemed at local shops. Comparison households could buy WaterGuard through normal retail channels, at about $0.25 for a 1-month supply (roughly a quarter of the agricultural daily wage). Although 70–90% of households in the study region had heard of the local brand of point-of-use chlorine and roughly 70% volunteered that drinking dirty water is a cause of diarrhea, only 5–10% of households reported that their main supply of drinking water was chlorinated prior to the interventions. There were no significant differences between treatment and comparison groups at baseline. Access to free chlorine increased takeup rates to over 50%, whereas coupons for even a 50% discount hardly affected takeup relative to the comparison group. The point estimate suggests a four percentage point increase relative to the comparison group, but this is not statistically significant. This is evidence for very price elastic demand. Households with young children did not behave differently from other households (p value of 0.85 on the test of equality of means). This evidence for low valuation of child health can be part of a distributional argument for subsidizing water treatment. Preliminary results from DeVoto et al. (2009) on distribution of chlorine through clinics in Kenya and from Berry et al. (2008) on distribution of water filters in Ghana also suggest very steep demand curves for improved water quality. The authors conclude that high takeup rates can be achieved at sufficiently low prices, but that demand for chlorine among their rural samples is extremely sensitive to price even though the retail price of the product is still relatively low. Ashraf et al. (2009) used a two-stage price randomization that enables both measurement of willingness to pay for water treatment and also, under specific assumptions, allows for testing of whether higher prices induce a sunk-cost effect that

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leads households who pay more for chlorine to use it more and/or screens out households less likely to use the product. In a door-to-door marketing campaign, roughly 1000 households in the study were first asked if they wanted to purchase a bottle of dilute chlorine at a randomized offer price. If a household agreed to purchase and was able to come up with the cash needed for the transaction, they were then offered an additional randomly assigned discount which determined the transaction price. Variation in offer prices is used to test for whether households who are willing to pay a higher price are more likely to use chlorine for water treatment, controlling for the price the household ultimately did pay; variation in transaction prices is used to test for a sunk-cost effect that might lead households who actually paid more for chlorine to be more likely to use it, controlling for willingness to pay. Approximately 2 weeks after the marketing campaign, the survey team reached almost 900 of the households who received the marketing intervention to test for the presence of chlorine in stored drinking water supplies and administered a follow-up study. The authors found that willingness to pay is low, and that many more households are willing to purchase chlorine at low prices. Consistent with other evaluations, Ashraf et al. did not find that charging higher price leads to more effective targeting to those households with higher potential health gains, those households with children under age 5 or pregnant women. There are at least two possible counter-arguments to the case for subsidies in water quality: (1) a sunk-cost effect, in which paying for the product makes people more likely to use it and (2) wastage of the product when people are given it but do not use it or use it for purposes not valued by the funder. Ashraf et al. (2009) found no evidence of a sunk-cost effect, finding that the actual transaction price does not affect propensity to use, controlling for offer price. Ashraf et al. (2009) did find that when the price is lowered, the marginal households induced to buy chlorine are less likely to show chlorine residual in their water 2 weeks later. The hypothesis that they favor is that these households start using the products for other off-label uses such as cleaning clothes or toilets. They presented evidence for this hypothesis drawn from a convenience sample. However, this is somewhat puzzling since, as they note, dilute chlorine sold for water treatment is considerably more expensive per unit of chlorine than commercially available bleach. Overall, it is difficult to distinguish the hypothesis that these households are not using the product for water treatment from the alternative hypotheses that they are storing the product or giving it away for water-treatment usage (or that they tried the product but did not like the taste; this would be only a short-run loss). These hypotheses have quite different policy implications as only the first is wastage that might reduce the social value of a program that supplied the product for free. Even if diversion to alternative uses is common, because chlorine is very cheap and can have a large impact on health, high levels of diversion to alternate uses are likely to be acceptable if this occurs as a result of a process that increases use for water treatment overall. Policymakers confronted with this evidence of low valuation of water quality and child health must determine whether subsidies for water-quality interventions such as chlorination are warranted. If governments or external donors

place more value on child health relative to other consumption compared to local households, the lack of valuation for water quality and child health provides a potential rationale for subsidies. Externalities from consumption provide another potential rationale for subsidies in some cases. Although there is no direct evidence on health externalities from water treatment in any of the papers reviewed here, to the extent that consumption reduces disease incidence for the user, it is also likely to reduce disease transmission from the user to others. In this case, eliminating prices for water-quality improvements is likely to be welfare maximizing due to these externalities. In fact, given the externalities combined with the low cost of water disinfectant, negative prices may be optimal.

1.05.4 Nonprice Determinants of Clean Water Adoption In this section, we review experimental evidence on several nonprice variables that could potentially affect household behavior regarding water quality. The emphasis in this section is on identifying potential mechanisms that could increase uptake of safe water rather than also judging their cost effectiveness or scalability. Section 1.05.5 discusses potentially scalable models drawing on the lessons of this and previous sections.

1.05.4.1 Information on Water Contamination Levels Several papers suggest that providing households with information about source water quality can change behavior, but that the effects of information are small relative to price and that people are responding not as Bayesian decision makers rationally processing information. Rather, information may be important because it increases the salience of water contamination. Jalan and Somanathan (2008) randomly assigned households in their urban Indian sample to receive information on whether or not their drinking water had tested positive for fecal contamination. Among households not purifying their water initially, this information led to an 11 percentage point increase in reported water purification as measured 8 weeks after information provision; about 42% of the study population purified (meaning either filtered, boiled, purchased bottled water, or, more rarely, chemically treated) their water at baseline. They also increased water purification expenditures by about $7. Households that initially purified their water but who received information that their water was probably not contaminated (based on tests of untreated water in their household) were not statistically more likely to change their purification behavior than the control group. This finding of an asymmetric response to testing is evidence for the idea that the channel through which information campaigns work is salience of some sort rather than Bayesian learning. Bayesian learners would respond to information that their water is safer or cleaner than they thought by reducing expenditure on purification. Luoto (2009) found through a randomized controlled trial in Kenya that sharing information on fecal contamination

Providing Clean Water: Evidence from Randomized Evaluations

with Kenyan households in a context in which treatment products were provided for free increases water treatment by 8–13 percentage points (or between 12% and 23% of baseline usage rates). The study also suggests that once information on the quality of source water is provided, providing additional information on the quality of water stored in the home has no further impact on takeup. When interpreting these results, it is useful to recall that, if people had unbiased expectations about water quality initially, information provision would lead some to revise beliefs about water quality downward and others to revise them upward, with ambiguous implications for water-treatment behavior. The fairly consistent finding that information provision increases takeup suggests that these results may reflect salience as much as Bayesian learning. Further evidence consistent with this is provided by Madajewicz et al. (2007) and Tarozzi et al. (2009), who studied as to how people respond to information about water quality in an area of Bangladesh where wells are frequently contaminated with arsenic. Madajewicz et al. evaluated the effectiveness of providing coarse information about well safety by providing a random sample of household information about whether their water source has arsenic concentrations above a threshold level. Households informed that their water exceeds this threshold are 37 percentage points more likely to switch sources than control households within 1 year. They increase their walking time 15-fold (about 4 min), on average, in response to the information. These responses to arsenic contamination information are further evidence for non-Bayesian decision making when combined with more recent research from Bangladesh following the introduction of a standardized labeling system for wells in which safe sources are labeled green and unsafe sources red. Tarozzi et al. (2009) performed an evaluation in this context in which all subjects receive the coarse information about water safety for all sources around them. A random sub-sample receives additional information about relative safety along a continuous scale. The relationship between arsenic and health is likely to be continuous. Thus, if households are Bayesian decision makers, continuous information should be more useful. Households far from any nearby uncontaminated well might switch to a well that is just above the cutoff level for being colored red, for example. Similarly, households using a well that is just below the cutoff might switch to a well that is much further below the cutoff. In practice, however, receiving continuous information does not substantially affect risk perceptions or the likelihood of switching sources. In fact, providing continuous information decreases the impact of the arsenic level on the probability of switching to a new source of drinking water. People are unable to use this additional information to improve their drinking water quality more than those people armed with only coarse information. This finding that coarse information may change behavior more than finer information poses a challenge to the idea that people can be modeled as Bayesian decision makers. These findings are instead consistent with the idea that information campaigns increase salience of water contamination and that it is this salience, which changes behavior rather than the precise information content itself. Similarly, the finding of an

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asymmetric response to information, with those who find that their water is contaminated adopting safer behavior, and those with uncontaminated water not reducing their efforts to obtain clean water, is unexpected under a Bayesian model, but is consistent with the salience hypothesis.

1.05.4.2 Gain versus Loss Framing and Other Behavioral Marketing Given the evidence above that a simple Bayesian learning story is inadequate to explain behavior, we now turn to evidence on ideas from behavioral economics and psychology. Luoto (2009) provided households a variety of point-of-use watertreatment technologies for free in Kenya and then randomly assigned households to receive various promotional strategies to increase use of these products. This research generates a series of results that can inform marketing and distributional strategies. First, she examines whether emphasizing the gains from water treatment versus the losses from not treating water affected use. There are competing hypotheses in the literature for which framing should bring about the larger response. Prospect theory predicts that loss aversion will cause the lossframed message to realize a bigger effect on people’s choices and behavior (Tversky and Kahneman, 1981; Kahneman and Tversky, 1979). However, there is evidence that decisions regarding health behaviors respond more to gain-framed messages in some cases and more to loss framings in others (Rothman et al., 1999). The study compares a framing of safe water technologies as increasing health compared to one in which it is framed as both increasing health and avoiding disease. The latter approach increased usage by approximately four to six percentage points, a statistically significant difference. Luoto (2009) also tested whether a combination of commitment and a visual reminder to treat water changes behavior. A subset of the sample was assigned to make a commitment to treating their water to improving their family’s health, and also given a pictorial reminder to treat their water. This increased water treatment by five to eight percentage points, but was significant only in some specifications. A commitment to the interviewer had relatively large effects on households that showed evidence of high discount rates in responses to hypothetical questions about future payoffs.

1.05.4.3 Communal versus Individual Persuasion Kremer et al. (2009a) provided some evidence that a communal approach in which households are aware of the messages other community members receive is more effective than an individual approach in encouraging treatment of household drinking water with dilute chlorine disinfectant, although differences are limited to the case when households had to pay for the product. Their study tested three variants of a persuasion campaign in which promotional messages targeted to mother were delivered at either the household level or community level, or both. The treatment was cross-cut with providing subsidized (free) chlorine to households. The results confirm the importance of price as a key determinant of takeup. When chlorine was subsidized, community messages had no measurable impact on household

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water treatment. The point estimate of the effect of messaging is actually negative, though statistically insignificant, and very small compared to the main effect (–0.02 as compared to 0.52). Messages can influence takeup, however, when positive prices are charged. At normal retail prices, treatment of household drinking water with chlorine increased by between three and five percentage points (as measured by testing household drinking water for chlorine) for the communitybased and combined scripts in the short run. There was no measurable impact of the household script alone, but community-based messaging, a much cheaper approach to marketing, had a small but positive effect. None of the promotion scripts had any significant effect on takeup at the medium-run follow-up 3–6 months after exposure. Considering the shortrun nature of the effects and the high cost of marketing during one-on-one conversations during household visits, or even through community-level meetings, such strategies do not appear to hold much promise as cost-effective means of promoting individually packaged retail chlorine takeup at scale. It is worth noting, however, that Kremer et al. (2009a) found little evidence for peer effects in takeup of chlorine packaged for household use. Using detailed data on conversation frequency and topics collected in the second and fourth survey rounds (of the first phase of the research), they found strong evidence that the distribution of free chlorine marketed as WaterGuard promoted conversations about the product as well as about drinking water more generally and, to a lesser degree, child health. In particular, conversations about WaterGuard were roughly 3 times more likely to occur if the respondent was a member of a treatment household and slightly more than twice as likely if the other household in a relationship pair was in the treatment group. Although the distribution of free WaterGuard prompted more conversations about the product, the evidence is consistent with the hypothesis of weak social network effects on actual use, with larger impacts on social desirability bias. They found statistically and economically significant effects of peer exposure on self-reported chlorination but pointed estimates that are much smaller and not generally statistically significant using positive chlorine tests in home drinking water.

1.05.4.4 Personal Contact Kremer et al. (2009a) also tested another approach to increasing demand – hiring local community members to promote chlorine use among their neighbors. This sort of personal contact has been previously identified as important to behavior change and adoption decisions (DellaVigna and Gentzkow, 2010; Manandhar et al., 2004). In this intervention, personal persuasion was accompanied by a price discount, however. Households were also given a coupon for one free bottle of dilute chlorine solution (equivalent to 1 month’s supply). The fraction of households with residual chlorine in their water was approximately 10 times as high in communities with a local promoter and a free sample of chlorine relative to comparison households in the short-run (3 weeks), at 40% roughly versus 4%, respectively. While takeup fell to a certain extent at the medium-run (3–6 months) as households used up their free bottle, communities with promoters were nonetheless able to

sustain adoption rates between 30 and 35 percentage points higher than the comparison group takeup rate of 8%. Personal contact was also successful in achieving high levels of takeup in the evaluation of the water-quantity intervention described in Section 1.05.2.1 (DeVoto et al., 2009).

1.05.5 Potentially Scalable Approaches to Improving Water Quality Section 1.05.2 documents the strong evidence that water treatment has the potential to improve health cost-effectively. Section 1.05.3 discusses that takeup of water quality interventions is extremely sensitive to price. Section 1.05.4 indicates that personal contact, salience, and convenience, and potentially having public information about water treatment can boost takeup. This section discusses potential low-cost, scalable models for water treatment based on the findings of Sections 1.05.2, 1.05.3, and 1.05.4. Kremer et al. (2009a) and DeVoto et al. (2009) developed and tested two alternative approaches to providing clean water, both involving free distribution. DeVoto et al. (2009) provided coupons for dilute chlorine solution to mothers who bring children to vaccination clinics sufficient to cover water supplies for the 12 months until children reach approximately age 2. Mothers are told how and where to redeem coupons and urged to treat water for their children during a vulnerable stage of their life. A second approach entailed switching to free delivery of chlorine by placing a container to dispense the product at water sources. This bulk supply dramatically reduced delivery costs relative to the retail approach that requires packaging chlorine in small bottles and makes free provision more realistic. Users can treat drinking water when they collect it. The required agitation and wait-time for chlorine-treated water are at least partially accomplished automatically during the walk home from the source. The source-based dilute chlorine disinfection approach to water treatment makes this act salient, convenient, and public, in addition to making it cheaper. The dispenser provides a daily visual reminder to households to treat their water at the moment when it is most salient – as water is collected – and maximizes the potential for learning, norm formation, and social network effects by making the dispenser public. Potential users can see others, who use the dispenser, and have the opportunity to ask questions and they will also know that others will see whether they use the dispenser. Takeup of chlorine provided through dispensers dramatically exceeded takeup of chlorine provided for in-home use. When communities were randomly assigned to treatment with a promoter and a community dispenser, takeup was about 40% in the short run (3 weeks) but had climbed to over 60% by the medium term (3–6 months), representing 37 and 53 percentage point gains respectively over the control group. In contrast to the takeup levels achieved with the dispensers, the clinic-based coupon distribution approach proved initially promising, but resulted in much lower coupon redemption over time. Over 40% of households, who were given coupons, redeemed them 8 months into the program in that sample, but this fell to 20% by 12 months. This suggests that the

Providing Clean Water: Evidence from Randomized Evaluations

success of the dispenser is not due only to the zero price, but also to the reduction in the psychic cost of remembering to treat water that is achieved by source-based treatment as well as other attributes, like the visual reminders. The chlorine dispenser is also extremely cost effective, with a cost per DALY saved of less than $20. The success of the chlorine dispensers at the proof-of-concept stage described here suggests that exploring how to scale up this approach to water treatment warrants further attention in its own right, as well. An important challenge for the future will be to determine how best to handle the supply side under free provision and, in particular, scaling supply chain management.

1.05.6 Methods and Theory: Contributions of Randomized Evaluations of Domestic Water The evaluations surveyed in this chapter have provided policy guidance on several questions related to health, technology adoption, and pricing regimes. The work has also made a number of methodological contributions that are of broader interest in resource economics. We review these contributions in this section.

1.05.6.1 Survey Effects A recent randomized evaluation of a water-quality intervention provides evidence that the act of surveying can affect behavior in ways that can interfere with estimates of treatment effects, a result with broader implications. Many studies have focused on measuring reported diarrhea (by mothers of young children) through household visits as a means of assessing the health impact of domestic water interventions. Kremer et al. (2009b) provided evidence that collection of self-reported diarrhea data through repeated interviews leads to health-protective behavior change in addition to respondent fatigue and social desirability bias, which are well-known concerns in survey administration. More frequent questioning exacerbates this survey effect. As part of a larger study of the impact of spring protection, households from a rural Kenyan population were randomly assigned to be interviewed about diarrhea either every 2 weeks or every 6 months. Kremer et al. (2009b) documented that frequent data collection, as is typically used in the epidemiology literature to measure diarrhea incidence, induces behavioral change in the form of higher levels of home water treatment, verified by tests for chlorine in water. The authors also find that frequent data collection leads to lower reports of child diarrhea by mothers relative to infrequent surveying. These effects are sufficiently large so as to change the conclusions about the effectiveness of the water-quality intervention being studied, that is, spring protection. The potential for survey effects implies that researchers relying on both self-reported or otherwise subjective data and objective data to measure outcomes should consider designing data collection strategies that minimize interaction with subjects. For example, outcome data could also be collected via administrative records maintained at clinics or schools. Purchases or collection of products from central locations can also be tracked without direct interaction with subjects.

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In the particular case of the literature on water, sanitation, and hygiene, survey effect concerns imply that more research is needed that does not measure impacts via subjective reports of diarrhea. Researchers in this field should expand their datacollection strategies to emphasize other health outcomes that can be measured objectively and infrequently. This will likely require larger sample sizes to detect small treatment effects (e.g., on stunting, cognition, and ultimately, mortality) and longer study times, which funding will need to accommodate.

1.05.6.2 Valuation: Revealed Preference versus Contingent Valuation One common approach to understanding willingness to pay for environmental amenities is to use stated preference data from hypothetical situations to identify the price that households would be willing to pay. The survey-based approach for eliciting stated preference is a method known as contingent valuation (CV), the development of which is surveyed by Hanemann (1994). While data from such studies may be the most practical solution when estimating valuation of a nonrival good or one that is costly to offer in a real transaction, it is also subject to a number of pitfalls stemming from the fact that choices in hypothetical situations might not be the same as those that the respondent would make in the real world, facing real budget constraints and real benefits (Diamond and Hausman, 1994). In addition, it may be difficult for individuals to know beforehand how they will value a good and survey respondents may strategically misstate their willingness to pay (Whittington, 2002). Another approach is estimating valuation of goods not traded in markets to use discrete choice models to analyze nonexperimental survey data on households’ decisions. (There are a number of papers, including Mu et al. (1990) that estimate more general demand functions for water from various sources. We do not focus on these papers here, since they do not explicitly deal with water quality as a measurable attribute in the decision process.) While this has the advantage of evaluating real choices, and hence providing information on a part of the household preference function (McConnell and Rosado, 2000), a potential problem with this type of method is that there may be unobservable household characteristics correlated with households’ choices as well as supplier pricing decisions, which will lead to biased results. For example, suppliers may charge higher prices where demand is higher, resulting in a positive cross-sectional correlation in prices and adoption even if increasing prices would result in lower demand for each household. A third approach is to infer willingness to pay by randomizing prices or locations of new facilities, thus generating random variation in travel time. This enables analysis based on actual choices while also addressing omitted variable bias, addressing the main concerns with both CV data and results based on nonexperimental data. This method also has the potential to enable examination of the allocative role of prices in targeting populations of interest and the isolation of specific channels of causality for effects of prices on demand. The literature on domestic water that we review here has provided one of the first direct comparisons of revealed preference and stated preference valuations of an environmental

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service in developing countries. Kremer et al. (unpublished manuscript) found that the stated preference approaches generate much higher valuation estimates than the revealed preference approach, by a factor of 3, with the contingent valuation survey approach exhibiting much greater dispersion, as well as considerable sensitivity to question framing, casting doubt on the reliability of stated preference methods. A revealed preference estimate of households’ valuation of cleaner water from source water-quality improvements can be assessed using a travel-cost approach that measures the number of trips made to the improved source relative to an unimproved source at a different distance, as long as source water-quality improvements are randomly assigned. Kremer et al. (unpublished manuscript) used such a method to estimate a model of demand for clean water from source quality improvements. As previously described, they exploited exogenous changes in the trade-off that households face when choosing between multiple water sources, some of which are close but contaminated and others of which are far but clean. They then contrasted this revealed preference estimate of willingness to pay for spring protection with two different stated preference methodologies: stated ranking of alternative water sources and contingent valuation. The divergence between the valuation estimates suggests that CV studies overestimate how much users value water-quality improvements.

1.05.6.3 Combining Randomized Evaluations with Structural Modeling Several recent papers combine data from randomized experiments with structural econometric methods in development economics (e.g., Todd and Wolpin, 2006). Kremer et al. (unpublished manuscript) combined experimental results with a structural model of water infrastructure investment to explore the implications of alternative property rights institutions on social welfare and assess the welfare impacts of alternative institutions governing water property rights. Using the valuation results discussed earlier as inputs into policy simulations, the authors can compare the welfare impacts of a number of potential scenarios under alternative sets of social norms regarding property rights. A hypothetical case of pure privatization, for example, in which landowners could restrict access to the spring and charge for water, results in relatively little investment in environmental protection (i.e., spring protection) since households’ willingness to pay for cleaner water is low, but leads to large static losses since landowners can extract consumer surplus by charging for even unprotected spring water even though the marginal cost of provision is zero. They concluded that, at low income levels, common property likely yields greater social welfare than private property, but that at higher income levels private property may yield higher social welfare.

1.05.7 Conclusion As noted in the introduction, the sole quantifiable environmental goal selected by the United Nations as part of the United Nations Millennium Development Goals is to reduce by half the proportion of people without sustainable access to

safe drinking water. Standard public finance theory suggests that local governments may ordinarily be best placed to allocate funds among competing local public goods, but it also suggests that international donors and national governments could reasonably intervene to support water projects due either to disease externalities across political jurisdictions, or to different distributional preferences than local decision makers. Both factors point to an emphasis on the impact of water on communicable diseases of children. While there is currently limited evidence on the health impact of increasing access to water without improving quality, there is strong evidence that improving water quality has health benefits for young children. Since households seem more willing to pay for access to increased quantities of water than for safer water, and since investments in water treatment are extremely cost-effective relative to other health expenditures, even expenditures such as vaccination, there seems a strong case for zero prices or even negative prices for water treatment. This chapter reviews evidence from randomized evaluations that can inform this policy debate on the quality versus quantity investment decision, and strategies to drive takeup of water treatment products. One promising approach is chlorine dispensers. Takeup of chlorination via communal chlorine dispensers (Kremer et al., 2009a), combined with a local promoter, is between 60% and 70% and the authors estimate that the long-run cost of supplying a community with bulk chlorine through a dispenser is only about 1/4 to 1/3 as much as with individually packaged bottles. This amounts to a cost of about 15 cents per person per year. The feasibility of this approach depends on the ability to solve the challenge of refill servicing and efforts to increase the density of dispensers (to drive refill costs down). Additional work to understand how to combine interventions and transition to greater levels of service as incomes rise remains an important area of policy-relevant work. The methodological lessons from the research on water-treatment takeup and valuation reviewed here can inform study design on the scale-up of alternative approaches to water-treatment and other experiments in resource economics, as well.

References Arnold B and Colford J (2007) Treating water with chlorine at point-of-use to improve water quality and reduce diarrhea in developing countries: A systematic review and meta-analysis. American Journal of Tropical Medicine and Hygiene 76(2): 354--364. Ashraf N, Berry J, and Shapiro J (in press) Can higher prices stimulate product use? Evidence from a field experiment in Zambia. American Economic Review. Banerjee A, Cole S, Duflo E, and Linden L (2007) Remedying education: Evidence from two randomized experiments in India. Quarterly Journal of Economics 122(3): 1235--1264. Banerjee A and Duflo E (2009) Experimental approach to development. Annual Review of Economics 1: 151--178. Bennear LS and Coglianese C (2005) Measuring progress: Program evaluation of environmental policies. Environment: Science and Policy for Sustainable Development 47(2): 22--39. Berry J, Fischer G, and Guiteras R (2008) Willingness to pay for clean water. Presented at Workshop on Scaling up Distribution of Water Treatment Technologies in Developing Countries, Harvard University, Cambridge, MA. Cardenas JC (2009) Experiments in environment and development. Annual Review of Resource Economics 1: 157--182.

Providing Clean Water: Evidence from Randomized Evaluations

Chattopadhyay R and Duflo E (2004) Women as policy makers: Evidence from a randomized policy experiment in India. Econometrica 72(5): 1409--1443. Clasen T, Brown J, Collin S, Suntura O, and Cairncross S (2004) Reducing diarrhea through the use of household-based ceramic water filters: A randomized, controlled trial in rural Bolivia. American Journal of Tropical Medicine and Hygiene 70(6): 651--657. Clasen T, Roberts I, Rabie T, Schmidt W, and Cairncross S (2006) Interventions to improve water quality for preventing diarrhoea. Cochrane Database of Systematic Reviews 2006, Issue 3. Art. No.: CD004794. DOI: 10.1002/14651858.CD004794. pub2. Conley T and Udry C (2001) Social learning through networks: The adoption of new agricultural technologies in Ghana. American Journal of Agricultural Economics 83(3): 668--673. Curtis V, Cairncross S, and Yonli R (2000) Domestic hygiene and diarrhoea – pinpointing the problem. Tropical Medicine and International Health 5(1): 22--32. Cutler D and Miller G (2005) The role of public health improvements in health advances: The 20th century United States. Demography 42(1): 1--22. DellaVigna S and Gentzkow M (2010) Persuasion: Empirical evidence. Annual Review of Economics 2 (doi:0.1146/annurev.economics.102308.124309). DeVoto F, Duflo E, Dupas P, Pariente W, and Pons V (2009) Happiness on tap: The demand for and impact of piped water in urban Morocco. Working Paper UCLA. Diamond P and Hausman J (1994) Contingent valuation: Is some number better than no number? Journal of Economic Perspectives 8(4): 45--64. Duflo E, Dupas P, and Kremer M (2007) Peer effects, pupil–teacher ratios, and teacher incentives: Evidence from a randomized evaluation in Kenya. Cambridge: Mimeo; Harvard University. Duflo E, Glennerster R, and Kremer M (2008a) Using randomization in development economics research: A toolkit. In: Shultz TP and Strauss J (eds.) Handbook of Development Economics, 4, pp. 1--2. Amsterdam: Elsevier. Duflo E and Kremer M (2008) Use of randomization in the evaluation of development effectiveness. Evaluating Development Effectiveness 7: 93--120. Duflo E, Kremer M, and Robinson J (2009) Nudging farmers to use fertilizer: Theory and experimental evidence from Kenya. Working Paper, 15131, NBER. Eckel C and Grossman P (1998) Are women less selfish than men? Evidence from dictator experiments. Economic Journal 108(448): 726--735. Esrey SA (1996) Water, waste, and well-being: A multicountry study. American Journal of Epidemiology 143(6): 608--623. Esrey SA, Potash JB, Roberts L, and Shiff C (1991) Effects of improved water supply and sanitation on ascariasis, diarrhoea, dracunculiasis, hookworm infection, schistosomiasis, and trachoma. Bulletin of the World Health Organization 69(5): 609--621. Fewtrell L, Kaufmann RB, Kay D, Enanoria W, Haller L, and Colford JM (2005) Water, sanitation, and hygiene interventions to reduce diarrhoea in less developed countries: A systematic review and meta-analysis. Lancet Infectious Disease 5: 42--52. Foster A and Rosenzweig M (2001) Imperfect commitment, altruism, and the family: Evidence from transfer behavior in low-income rural areas. Review of Economics and Statistics 83(3): 389--407. Galiani S, Gertler P, and Schargrodsky E (2005) Water for life: The impact of privatization of water services on child mortality. Journal of Political Economy 113(1): 83--119. Gamper-Rabindran S, Khan S, and Timmens C (2010) The impact of piped water provision on infant mortality in Brazil: A quantile panel-data approach. Journal of Development Economics 92(2): 188–200. Garrett V, Ogutu P, Mabonga P, et al. (2008) Diarrhoea prevention in a high-risk rural Kenyan population through point-of-use chlorination, safe water storage, sanitation, and rainwater harvesting. Epidemiology and Infection 136(11): 1463--1471. Greenstone M and Gayer T (2009) Quasi-experimental and experimental approaches to environmental economics. Journal of Environmental Economics and Management 57(1): 21--44. Han AM and Hlaing T (1989) Prevention of diarrhoea and dysentery by hand washing. Transactions of the Royal Society of Tropical Medicine and Hygiene 83(1): 2128--2131. Hanemann WM (1994) Valuing the environment through contingent valuation. Journal of Economic Perspectives 8: 19--43. Hoffman V (2009) Intrahousehold allocation of free and purchased mosquito nets. American Economic Review 99(2): 236--241. Holla A and Kremer M (2008) Pricing and access: Lessons from randomized evaluation in education and health. Cambridge: Mimeo; Harvard University. Imbens G and Wooldridge JM (2009) Recent developments in the econometrics of program evaluation. Journal of Economic Literature 47(1): 5--86.

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Jalan J and Somanathan E (2008) The importance of being informed: Experimental evidence on demand for environmental quality. Journal of Development Economics 87: 14--28. Kahneman D and Tversky A (1979) Prospect theory: An analysis of decision under risk. Econometrica 47(2): 263--291. Khan MU (1982) Interruption of shigellosis by hand washing. Transactions of the Royal Society of Tropical Medicine and Hygiene 76(2): 164--168. Kremer M, Leino J, Miguel E, and Zwane A (2008) Managing rural water infrastructure in Kenya. Working Paper. Kremer M and Miguel E (2007) The illusion of sustainability. Quarterly Journal of Economics 112(3): 1007--1065. Kremer M, Miguel E, Mullainathan S, Null C, and Zwane A (2009a) Coupons, promoters, and dispensers: Impact evaluations to increase water treatment. Working Paper. Kremer M, Miguel E, Null C, Van Dusen E, and Zwane A (2009b) Measuring diarrhea: Quantifying Hawthorne effects in frequently collected data. Working Paper, University of California, Berkeley. Laibson D (1997) Golden eggs and hyperbolic discounting. Quarterly Journal of Economics 112(2): 443--477. Lancet (2007) Science at WHO and UNICEF: The corrosion of trust. Lancet 370: 1007 (editorial). Lipscomb M and Mobarak M (2008) Decentralization and water pollution spillovers: Evidence from the re-drawing of county boundaries in Brazil. Working Paper. molly.lipscomb.googlepages.com/Brazilwater113007.pdf (accessed March 2010). Luby S, Agboatwalla M, Painter J, Altaf A, Billhimer W, and Hoekstra H (2004) Effect of intensive hand washing promotion on childhood diarrhea in high-risk communities in Pakistan: A randomized control trial. JAMA 291(21): 2547--2554. Luby S, Mendoza C, Keswick B, Chiller T, and Hoekstra R (2008) Difficulties in bringing point-of-use water treatment to scale in rural Guatemala. American Journal of Tropical Medicine and Hygiene 78(3): 382--387. Luoto J (2009) Information and persuasion: Achieving safe water behavior in Kenya. Working Paper. Madajewicz M, Pfaff A, van Geen A, et al. (2007) Can information alone change behavior? Response to arsenic contamination of groundwater in Bangladesh. Journal of Development Economics 84: 731--754. Manandhar DD, Osrin B, Prasad N, et al. (2004) Effect of a participatory intervention with women’s groups on birth outcomes in Nepal: Cluster randomized control trial. Lancet 364: 970--979. Mansuri G and Rao V (2004) Community-based and -driven development: A critical review. World Bank Research Observer 19(1): 1--39. McConnell K and Rosado M (2000) Valuing discrete improvements in drinking water quality through revealed preferences. Water Resources Research 36(6): 1575--1582. Miguel E and Gugerty M (2005) Ethnic diversity, social sanctions, and public goods in Kenya. Journal of Public Economics 89(11–12): 2325--2368. Mu X, Whittington D, and Briscoe J (1990) Modeling village water demand behavior: A discrete choice approach. Water Resources Research 26(4): 521--529. Mullainathan S, Schwartzstein J, and Shleifer A (2006) Coarse thinking and persuasion. Working Paper, W12720, NBER. Munshi K (2004) Social learning in a heterogeneous population: Technology diffusion in the Indian Green Revolution. Journal of Development Economics 73: 185--215. Muralidharan K and Sundararaman V (2009). Teaching incentives in developing countries: Experimental evidence from India. Working Paper, 15323, NBER. Nowell C and Tinkler S (1994) The influence of gender on the provision of a public good. Journal of Economic Behavior and Organization 25(1): 25--36. O’Donoghue T and Rabin M (1999) Doing it now or later. American Economic Review 89(1): 103--124. Oster SM (1995) Strategic Management for Nonprofit Organizations: Theory and Cases. Oxford: Oxford University Press. Pattanayak SK and Pfaff A (2009) Behavior, environment, and health in developing countries: Evaluation and valuation. Annual Review of Resource Economics 1: 183--217. Ray I (2004) Water for all? Peri-urban and rural water delivery options: The case of India. International Conference of Engineers for a Sustainable World, Stanford Unversity. Rothman A, Martino S, Bedell B, Detweiler J, and Salovey P (1999) The systematic influence of gain- and loss-framed messages on interest in and use of different types of health behavior. Personality and Social Psychology Bulletin 25(11): 1355--1369. Sachs JD (2005) The End of Poverty: Economic Possibilities for Our Time. New York: Penguin.

1.06 Pricing Water and Sanitation Services D Whittington, University of North Carolina at Chapel Hill, Chapel Hill, NC, USA & 2011 Elsevier B.V. All rights reserved.

1.06.1 1.06.2 1.06.3 1.06.4 1.06.5 1.06.5.1 1.06.5.1.1 1.06.5.1.2 1.06.5.1.3 1.06.5.1.4 1.06.5.1.5 1.06.5.2 1.06.5.3 1.06.6 1.06.7 1.06.7.1 1.06.7.2 1.06.8 References

Introduction The Costs of Providing W&S Services W&S Development Paths Objectives of Tariff Design Tariff Structures – the Alternatives Single-Part Tariffs Fixed charges Volumetric charges Uniform volumetric charge Block tariffs Increasing linear tariff Two-Part Tariffs Seasonal and Zonal Water Pricing Achieving Economic Efficiency and Recovering Capital Costs: Fundamentals of Dynamic Marginal Cost Pricing in the W&S Sector Subsidizing Capital Costs: Reaching the Poor Create a Well-Run System of Public Taps as a Safety Net for the Poor Preserve Options for the Poor Concluding Remarks

1.06.1 Introduction A tariff is an important management tool that can be used to assist with efforts to improve the delivery of water and sanitation (W&S) services. The pricing of W&S services is, however, controversial, and it is important to understand why there is so little consensus on W&S tariff issues. There are four main reasons. First, there is disagreement over the objectives of water pricing and tariff design. Water pricing decisions affect several different objectives or goals of policymakers, often in conflicting ways. This means that if one person is looking solely (or mostly) at the consequences of a particular water pricing policy (or tariff design) in terms of one objective, and another person is looking at the same water pricing policy in terms of its impact on another objective, they may reach quite different conclusions about the attractiveness of the pricing policy. Second, as people do not generally know what it costs to provide W&S services, it is difficult for them to judge what is a fair or appropriate price to pay. Third, there is disagreement over what would actually happen if different water tariffs were implemented. The empirical work is often lacking that would enable someone to know with reasonable confidence how changes in water prices would affect the quantity of water that different customers would use and whether or not price changes would affect customers’ decisions to connect (or stay connected) to the water distribution system (Nauges and Whittington, 2009). Fourth, although there is some competition in the water market, there is no market test for different water tariff structures. Many tariff structures are feasible and can partially accomplish some of the competing objectives of water pricing.

79 79 82 83 84 85 85 85 85 86 87 87 87 88 91 93 94 94 94

There are typically an insufficient number of providers of piped water services for customers to reject inappropriate tariff structures. Bad ideas thus do not get weeded out of either sector practice or policy discussions. Even in different private sector participation arrangements, water tariff structures are typically set by the regulatory agency, and the private sector operator has to treat them as given and manage the system as best he can (given this constraint). The purpose of this chapter is to provide the reader with a better understanding of the main issues involved in the design of W&S tariffs. Section 1.06.2 summarizes the costs of providing piped W&S services. Obviously, these costs vary widely depending on local circumstances, but the presentation of some estimates of different components of the costs of providing such services illustrates that they are not cheap. Section 1.06.3 discusses alternative development paths for moving from low levels of W&S service (or no service at all) to modern piped services, and shows the costs associated with various incremental changes. Section 1.06.4 presents the four main objectives of tariff design. Section 1.06.5 summarizes the main tariff options. Section 1.06.6 describes the basic ideas of dynamic marginal cost pricing in the W&S sector and illustrates how a two-part tariff can be used to achieve both economic efficiency and cost recovery objectives. Section 1.06.7 discusses how subsidies can be best used in the W&S sector to reach poor households. Section 1.06.8 offers some concluding remarks.

1.06.2 The Costs of Providing W&S Services A key feature of network W&S investments is that they are very capital intensive. The majority of costs are incurred early in the

79

80

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life of the project, and the subsequent stream of benefits (and revenues) occurs over many years. Baumann and Boland (1998) estimated that the ratio of annual investment in the US water industry to gross revenues was 0.43 in 1993. (They note, ‘‘no other major industry group in the United States even approaches this ratio of annual investment to revenue.’’) Someone (e.g., private investors and government) must take a long-term perspective and put large amounts of capital at risk. If future revenues are needed to pay for the high capital costs in the early years, investors need assurance to their rights to the revenue stream. In contrast, poor households in developing countries tend to have high rates of discount (Poulos and Whittington, 2000) and thus have short planning horizons. Such households cannot easily make the long-term commitments required to pay for network W&S services. They are also uncertain about the prospects for long-term economic growth and their ability to pay in the future. The challenge of W&S tariff design in developing countries must be understood within the context of this fundamental mismatch of perspectives between investors and consumers. In urban areas, there is widespread consensus that the long-term goal of W&S service providers in developing countries should be to offer 24-h potable water supply piped into people’s homes, to remove wastewater with a piped sewerage system, and to treat this wastewater to a standard sufficient to minimize the environmental effects of its discharge to surface water bodies. (Some people in the sector do question the wisdom of pursuing the goal of piped sewerage – see, e.g., Esrey and Andersson (2000).) Even in many rural communities, households aspire to this level of service. The treatment and delivery of water to households, and the removal and treatment of the wastewater generated, cost serious money. These costs must be paid by someone, households must either pay or receive subsidies (e.g., from richer households, industries, donors, and higher levels of government). Of course, costs vary depending on local circumstances, and estimates of what it will cost to provide a certain level of service may vary widely. Also, most investments are incremental in nature. Only rarely would a community incur the costs of complete (full service) piped W&S systems at a single point in time. Nevertheless, some rough calculations may prove useful for the discussion of water pricing and tariff design. The approach here is to present some illustrative average unit costs of providing an urban household with modern W&S services. First, representative unit costs per cubic meter for different components of W&S services are looked at. Second, some typical quantities of water that different representative households use in a month are provided. Third, representative unit costs are multiplied by typical monthly household water use to obtain estimates of the monthly economic costs of providing a typical household with improved, piped W&S services. The economic costs of providing a household with modern W&S services are the sum of seven principal components: 1. opportunity costs of diverting raw water from alternative uses to the household (or resource rents); 2. storage and transmission of untreated water to the urban area;

3. treatment of raw water to drinking water standards; 4. distribution of treated water within the urban area to the household; 5. collection of wastewater from the household (sewerage collection); 6. treatment of wastewater (sewage treatment); and 7. any remaining costs or damages imposed on others by the discharge of treated wastewater (negative externalities). Table 1 presents some illustrative average unit costs for each of these seven cost components, expressed in US$ per cubic meter. The unit costs of these different cost components could vary widely in different locations. For example, in a location with abundant freshwater supplies, the opportunity cost of diverting water from existing or future users to our illustrative household (item 1) and the damages imposed by the discharge of treated wastewater (item 7) may, in fact, be very low or even zero. However, in more and more places, these opportunity costs associated with water diversion and the externalities from wastewater discharge are beginning to loom large. Some cost components are subject to significant economies of scale, particularly storage and transmission (item 2), the treatment of raw water to drinking water standards (item 3), and the treatment of sewage (item 6). This means that the larger the quantity of water or wastewater treated, the lower the per-unit cost. On the other hand, some cost components are experiencing diseconomies of scale. As large cities go farther and farther away in search of additional freshwater supplies, and good reservoir sites become harder to find, the unit cost of storing and transporting raw water to a community increases. There are also trade-offs between different cost components: one can be reduced, but only at the expense of the other. For example, wastewater can receive only primary treatment, which is much cheaper than primary and secondary treatment, but then the negative externalities associated with wastewater discharge will increase. Table 1 Cost estimates: improved water and sanitation services (assuming 6% real discount rate, see Whittington et al. (2008) for details) No.

Cost component

US$/m3

1

Opportunity cost of raw water supply Storage and transmission to treatment plant Treatment to drinking water standards Distribution of water to households (including house connections) Collection of wastewater from home and conveyance to wastewater treatment plant Wastewater treatment Damages associated with discharge of treated wastewater

0.05

3

0.10

5

0.10

5

0.60

30

0.80

40

0.30 0.05

15 3

2.00

100

2 3 4

5

6 7 Total

% of total

Pricing Water and Sanitation Services

The cost estimates in Table 1 include both capital expenses, and operation and maintenance expenses. The opportunity costs of raw water supplies (item 1) are still quite low in most places, on the order of a few cents per cubic meter. Even in places where urban water supplies are diverted from irrigated agriculture, the unit costs will rarely be above US$0.25 per cubic meter. Desalinization and wastewater reclamation costs will set an upper limit on opportunity costs of raw water in the range of US$0.50–1.00 per cubic meter for cities near the ocean, but the opportunity costs of raw water are nowhere near this level in most places. Raw water storage and transmission and subsequent treatment (items 2 and 3) will typically cost US$0.20 per cubic meter. Within a city, the water distribution network and household connections to it (item 4) comprise a major cost component, in many cases on the order of US$0.60 per cubic meter. The collection and conveyance of sewage to a wastewater treatment plant (item 5) are even more expensive than the water distribution; this will cost about US$0.80 per cubic meter, 40% of the total cost. Secondary wastewater treatment (item 6) will cost about US$0.30 per cubic meter. Damages resulting from the discharge of treated wastewater are very sitespecific, but environmentalists correctly remind us that that they can be significant, even for discharges of wastewater receiving secondary treatment. Let us assume for purposes of illustration that these costs are of the same order of magnitude as the opportunity costs of raw water supplies (US$0.05). As shown, total economic costs are about US$2.00 per cubic meter in many locations. It is emphasized that costs shown here are not intended to represent an upper bound. For example, in small communities in the arid areas of the western United States’ costs of W&S services can easily be double or triple these amounts per cubic meter. Also, note that these cost estimates assume that financing is available at competitive international market rates, and that countries do not pay a high default or risk premium. Table 2 presents a reasonable lower-bound estimate of unit costs of piped W&S services. Here, the opportunity costs of raw water supplies and the damages from wastewater discharges are assumed to be zero. Only minimal storage is included, and the only intake treatment is simple chlorination. Costs for the water distribution network assume the use of polyvinyl chloride pipes and shallow excavation. Wastewater is collected with condominial sewers, and the only wastewater treatment is provided by simple lagoons. Given all these assumptions, one can manage to reduce unit costs of piped W&S services to about US$0.80 per cubic meter. How much water does a typical household in a developing country need? The quantity of water used by a household will be a function of the price charged, household income, and other factors. Currently, most households in developing countries are facing quite low prices for piped W&S services. One can look at water use data from households in industrialized countries to see how much water one might expect a household to use for a comfortable modern lifestyle. For households with an in-house piped water connection, in many locations residential indoor water use falls in the range of 110–220 l per capita per day. For a household of six members, this would amount to about 20–40 cubic meters per month (Table 3). At the current low prices prevailing in many

81

Table 2 Cost estimates: improved water and sanitation services for low-cost option for private water and sewer connections (assuming 6% real discount rate, see Whittington et al. (2008) for details) No.

Cost component

US$/m3

1

Opportunity cost of raw water supply (steal it) Storage and transmission to treatment plant (minimal storage) Treatment of to drinking water standards (simple chlorination) Distribution of water to households (PVC pipe) Collection of wastewater from home and conveyance to wastewater treatment plant (condominial sewers) Wastewater treatment (simple lagoon) Damages associated with discharge of treated wastewater (someone else’s problem)

0.00

2 3 4 5

6 7

Total

0.07 0.04 0.24 0.30

0.15 0.00

0.80

Table 3 Range of estimates of monthly water use (in-house, private connection) Per capita daily water use (l)

Persons per household

Days per month

Monthly household water use (m3)

55 110 220

6 6 6

30 30 30

10 20 40

cities in developing countries, such levels of household water use are not uncommon. Other things being equal, households living in hot, tropical climates use more water for drinking, bathing, and washing than households in temperate or cold climates. Assuming average unit costs of US$2.00 per cubic meter, the full economic costs of providing 20–40 cubic meters of water to a household (and then dealing with the wastewater) would be US$40.00–80.00 per month (Table 4), more than most households in industrialized countries pay for the same services and far beyond the means of most households in developing countries. One would expect poor households in developing countries with in-house water connections to respond negatively to higher W&S prices: they might curtail use to as little as 50–60 l per capita per day. For a household with six members, at 55 l per capita per day, total consumption would then amount to about 10 cubic meters per month. The full economic costs of this level of W&S service at this reduced quantity of water use (assuming our unit costs of US$2.00 per cubic meter remained unchanged) would then be US$20.00 per month per household. At entirely plausible levels of water use (110 l per capita per day), the total economic cost would be about US$40.00 per month for the same household. With the unit costs of the low-cost system depicted in Table 2, the full economic cost of providing 10 cubic meters per month would be US$8.00 per household per month. This estimate should be regarded as a

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Table 4 Range of estimates of the full economic cost of providing improved W&S services (in-house, private water connection; piped sewer) Monthly household water use (m3)

Average cost ¼ US$0.80/m3 (US$)

Average cost ¼ US$2.00/m3 (US$)

10 20 40

8 16 32

20 40 80

lower bound on the full economic costs of piped W&S services in most locations. In industrialized and developing countries alike, most people are unaware of the magnitude of the true economic costs of municipal W&S services. There are several reasons why these economic costs are so poorly understood. First, the capital costs are heavily subsidized by higher levels of government (and, in developing countries, by international donors), so that households with services do not see the true capital costs reflected in the fixed charges or volumetric prices they pay. Second, in many cities tariff structures are designed so that industrial water usage subsidizes residential usage; households thus do not even see the full operation and maintenance costs in the prices they pay. Third, as many water utilities run financial deficits (in effect running down the value of their capital stock), water users in aggregate do not even see the full costs of supply. Fourth, most cities do not pay for their raw water supplies: typically, the water is simply expropriated from any existing water sources (and their users) in outlying rural areas. Fifth, wastewater externalities are typically imposed on others (downstream) without compensation. Sixth, the subsidies provided to consumers of W&S services are not only huge, but also regressive. It is often not politically desirable for the majority of people to understand that middle- and upper-income households, who generally use more water, are thus actually receiving the most benefit from subsidies. Tariff designs may in fact be made overly complicated in order to offset this reality and appear to be helping poorer households (Komives et al., 2005). Most fundamentally, poor households are often not connected to the W&S network at all and hence cannot receive the subsidized services. Even if they do have connections, the poor use less water than richer households, thus receiving lower absolute amounts of subsidies. The estimates presented here are intended merely to suggest what W&S costs are like in many developing countries. A reasonable question to ask is whether costs differ much across countries in the developing world and between industrialized and developing countries. Labor costs are obviously lower in developing countries, but because W&S projects are capital intensive, the cost component has less of an impact on total costs than for other goods and services. There are, to our knowledge, no publicly available international indices of W&S project construction costs. To illustrate the magnitude of international cost differentials for some related goods and construction costs, Table 5 compares costs of rebar, cement, and industrial construction in 11 large cities in both

Table 5 Comparison of costs of rebar, cement, and industrial facility construction in 11 cities City

Rebar (US$/ton)

Cement (US$/ton)

Industrial Construction (US$/m2)

London Boston Los Angeles Shanghai Jakarta Bangkok Hanoi New Delhi Durban Nairobi Buenos Aires

981 1100 992 435 528 482 349 600 1028 NA 765

96 85 135 43 68 63 62 64 137 NA 82

850 915 699 592 269 301 409 247 516 291 NA

From Engineering News-Record (2004) 253. 24 (December 12), 32–37.

industrialized and developing countries. Costs are indeed lower in cities such as New Delhi and Hanoi than in London and Boston, and lower costs for inputs such as cement and steel will translate into lower costs for W&S projects. It is, of course, less expensive to provide intermediate levels of W&S services (e.g., public taps and communal sanitation facilities) than the costs in Table 2 would indicate. Monthly household costs for such services are, however, often quite considerable, roughly US$5.00–10.00 per month for much smaller quantities of water and much lower levels of sanitation services. These costs are often reported to be as low as US$1.00–2.00 per household per month, but such accounts often systematically underestimate key cost components and rarely reflect the real costs of financially sustainable systems. It is also important to appreciate that intermediate services impose additional costs on households in terms of extra time spent accessing the services and increased coping costs for the inconveniences of using off-site services (Bahl et al., 2004; Pattanayak et al., 2005).

1.06.3 W&S Development Paths The high capital costs of network water and sewer systems have important implications for water prices and tariff design. Decisions on how to price network W&S services in developing countries are typically made in a dynamic, changing environment. Pricing and tariff design decisions made today should not lock households into low-level equilibrium solutions that will constrain them from improving their W&S services as economic growth occurs. In-house piped W&S services are unaffordable today in many cities in developing countries, but as economic growth occurs, there is general agreement that this goal is both desirable and achievable. It is thus important to consider carefully how pricing and tariff-design decisions influence the evolution of W&S service provision and the ability of managers and planners to upgrade services when economic growth creates the resources to make this vision a reality. There are numerous strategies or development paths for moving from a situation where households have poor or no services to

Pricing Water and Sanitation Services

modern W&S services, and it is necessary to reflect explicitly on the pros and cons of each, and how pricing and tariff design decisions push service providers and households along a particular development path, or create hurdles that must be overcome to make progress. For purposes of illustration, Table 6 compares three levels of water services and four levels of household sanitation. Let us consider a household without either improved water or sanitation services (case 1). Within the parameters given in the table, such a household might progress from this status to full modern W&S services (case 12) along any of the four principal development paths. First, some water planners would advocate for a water-first development path (case 1 - case 2 - case 3 - case 6 - case 9 - case 12); here W&S service providers concentrate on first getting piped water services into the household; only after this stage is achieved would investments go to the installation of neighborhood sewers and then to wastewater treatment. Note that the household itself has important investments to make. On the water side, in-house plumbing is required to take full advantage of the piped water connection. Similarly, the household would typically be responsible for the installation of a private water-sealed toilet, without which the installation of neighborhood sewers would be of less value. Proponents of a water-first development path argue that people want water services first and do not recognize the need for removing wastewater from the household until water has been provided and wastewater removal has become a problem. Also, as described above, sewers and wastewater treatment are very expensive, so it is easier financially to provide the less-expensive services first. From a pricing perspective, a water-first strategy has important implications. Under this strategy, revenues from water sales should not be diverted to subsidize sewers or wastewater treatment, at least until the majority of the population has high-quality water services. Also, any available subsidies from

Table 6

83

higher levels of government should be used to push households toward in-house piped service. For example, subsidies might be used to reduce the upfront connection charge for a household water hookup. Public health professionals sometimes argue in favor of a second development path: not one without the other (case 1 - case 5 - case 9 - case 12). Proponents of this approach believe that there are important public health complementaries from providing improved W&S together, and that households should not be allowed to receive in-house piped water without hooking up to a sewer line. Engineers often point out that it is cheaper to install water and sewer lines at the same time, particularly in cities where this may entail tearing up streets, sidewalks, and other infrastructure. Such bundling of W&S services has important implications for tariff design. If households are required to have sewer services when they receive piped water services, then from a household’s point of view, W&S services cannot really be charged separately. (Also, note that W&S service providers cannot practically meter the amount of water that a household receives separately from the amount of wastewater that it discharges. Thus, even if a provider claims to calculate water and wastewater charges separately, and apply a separate volumetric charge to each flow, from the household’s perspective this is simply an accounting trick. The household effectively faces a single weighted volumetric rate for the combined service.) If the service provider attempts to recover the full costs of both services, and the household is willing to pay the cost of the water services but unwilling to pay for the sanitation services, the household will reject the entire bundle. Thus, when services are bundled and tariffs are designed to recover the costs of service, tariffs can easily become a barrier to the provision of full modern network services (case 12). A third development path might be termed ‘sanitation first’ (case 1 - case 4 - case 5 - case 8 - case 9 - case 12). The rationale here is that improved sanitation is a more important

Water and sanitation development paths Unimproved water source (e.g., pond and river)

Improved water source outside the home (e.g., hand-pump and public tap)

Improved water inside the home (private water connection or yard tap)

No improved sanitation

Case 1

Case 2 (US$5/month/household)

Case 3 (US$10/month/ household)a

On-site sanitation (e.g., VIP latrine and pour flush toilet)

Case 4 (US$5/month/household)

Case 5 (US$10/month/household)

Case 6 (US$15/month/household)

Water-sealed toilet þ neighborhood wastewater collection (e.g., small-bore or conventional sewers)

Case 7 (US$15/month/ household)b

Case 8 (US$20/month/household)

Case 9 (US$25/month/household)

Water-sealed toilet þ neighborhood wastewater collection þ wastewater treatment

Case 10 (US$25/month/ household)

Case 11 (US$30/month/ household)

Case 12 (US$35/month/ household)

a

Water costs are not cumulative because having a private connection does not require a public tap or handpump. Sanitation costs are cumulative, that is, level 3 includes the costs of in-house plumbing þ neighborhood wastewater collection.

b

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first step than improved water services in achieving the desired public health benefits. Thus, if resources are limited, public authorities should tackle sanitation problems before building piped water distribution networks. From a pricing perspective, if demand for improved sanitation services turns out to be low, this development path requires more initial subsidies, with revenues from water to follow when households eventually receive in-house water connections. A fourth set of development paths might be termed ‘demand-driven’ in that the paths are not selected by experts but rather by people themselves. There are numerous plausible development paths that households might choose (e.g., case 1 - case 3 - case 6 - case 9 - case 12; or case 1 - case 6 case 9 - case 12). If households’ preferences are allowed to shape the evolution of W&S services, prices and tariff design have an especially important role to play. Prices provide the signals about the real resource costs of the various steps from the status quo to full modern network services. If these signals are incorrect, households may take an unwanted or unnecessary detour on the road to case 12.

1.06.4 Objectives of Tariff Design Setting water (and sanitation) tariffs requires that one strikes a balance between four main objectives (see Boland (1993) and Whittington et al. (2002) for additional discussion of tariff objectives). Cost recovery. From the water supplier’s point of view, cost recovery is the main purpose of the tariff. (For example, the World Bank’s Operational Manual Statement No. 3.72 emphasizes the importance of the cost recovery objective and the financial autonomy of the borrower.) Cost recovery requires that, on aggregate, tariffs faced by consumers should produce revenue equal to the financial costs of supply. Moreover, the revenue stream should be relatively stable and not cause cash flow or financing difficulties for the utility. Economic efficiency. Economic efficiency requires that prices be set to ensure that customers face the avoidable costs of their decisions. In other words, prices should signal to consumers the financial, environmental, and other costs that their decisions to use water impose on the rest of the system and on the economy. In practice, this means that the volumetric charge should be set equal to the short-run marginal social cost of bringing one additional cubic meter of water into a city, delivering it to a particular customer, collecting and treating the wastewater, and discharging the treated wastewater into a receiving water body. In many cities, the cost of bringing in additional water is higher than the cost of supplying the water already on hand, as the cheapest sources tend to be developed first. The short-run marginal cost should include not only the financial cost of public works undertaken, but also the social cost of diverting water resources into public supply rather than using it for other purposes. An efficient tariff will create incentives that ensure, for a given water supply cost, that users obtain the largest possible aggregate economic benefits. Equity. The term ‘equity’ is often used to denote quite different things. Here, it is used to mean that the water tariff treats similar customers equally, and the customers in different

situations are not treated the same. This usually means that users pay monthly water bills that are proportionate to the costs they impose on the utility by their water use. Affordability. One objective of tariff design is to ensure that poor households are able to obtain adequate supplies of clean water. The terms ‘equity’, ‘fairness’, ‘poverty alleviation’, and ‘affordability’ are often used interchangeably to express this desire. It is preferred to treat affordability as an objective distinct from equity, fairness, and poverty alleviation, because a W&S tariff that is affordable may not be equitable or perceived as fair. Moreover, an affordable tariff may not pull poor households out of poverty. Many people feel that water services are a basic right and should be provided to people regardless of whether they can pay for the services. These considerations have led to recommendations that W&S tariffs should be kept low and that water should be provided free or at minimal cost, at least to the poor, through systems of subsidies. There are a number of trade-offs between these different objectives and the W&S tariffs used to calculate customers’ bills. For example, providing water free through private connections in order to achieve the objective of affordability conflicts with the objectives of cost recovery and efficient water use. Also, poor customers can sometimes be relatively expensive to serve (e.g., perhaps due to their outlying location), and hence it might not be regarded as equitable to charge them the same as, or less than, other customers. Additional objectives and considerations may be involved. For example, a tariff design should be easy to explain, understand, and implement. A tariff design should be acceptable both to the public and to the political leaders. This may require the tariff to conform to perceptions of fairness, often quite different from notions of equity. Water tariffs may be designed to discourage excessive uses of water, thus promoting water conservation, where excessive may be understood as a deviation from some notion of a fair amount. A successful W&S tariff design should not be controversial, nor should it become a focus of public criticism of the watersupply agency. Human beings are, however, acutely sensitive to situations perceived to be unfair, and fairness is often in the eye of the beholder. It can prove to be especially difficult to design a W&S tariff that is perceived to be fair when customers do not understand the true resource costs of providing modern W&S services. Consider the four cases in Table 7. If household members understand the real resource costs of supplying modern W&S services and believe the household should pay a share of these costs proportionate to its use of such services (case A), a W&S tariff that is perceived to be fair can be relatively easily designed. But if household members do not understand the real resource costs of supplying modern W&S services, it may prove to be difficult for them to believe that a tariff is fair even if they believe the household should pay a share of these costs proportionate to its use of such services (case B). For example, such a household may perceive a proposed W&S tariff to be price gouging even if it is not. On the other hand, household members may understand the real resource costs of supplying modern W&S services but not believe the household should pay a share of the costs proportionate to its use of such services (case C). This may be due to past injustices, a feeling that this household is more

Pricing Water and Sanitation Services Table 7 Households’ understanding of supply costs vs. agreement to pay a proportionate share of the costs of W&S services: four cases A household understands the real resource costs of supplying modern W& S services

A household does not understand the real resource costs of supplying modern W& S services

A household believes that it should pay a proportionate share of the costs of W& S services

Case A

Case B

A household believes that it should not pay a proportionate share of the costs of W& S services

Case C

Case D

deserving of help than others, or any number of reasons. Or household members may neither understand the real resource costs of supplying modern W&S services, nor believe the household should pay a share of the costs proportionate to its use of such services (case D). This is the most difficult situation for all stakeholders, and unfortunately it is quite common. In case C and especially in case D, the negotiation of W&S tariffs often becomes a political problem largely unrelated to the costs of service delivery.

1.06.5 Tariff Structures – the Alternatives A tariff structure is a set of procedural rules used to determine the conditions of service and the monthly bills for water users in various categories. (This section draws heavily on Whittington et al. (2002).) Table 8 presents a simple classification of the different types of water tariff structures. Two main types of tariff structures are used in the municipal water supply sector: a single-part tariff and a two-part tariff. With a singlepart tariff, a consumer’s monthly water bill is based on a single type of calculation. With a two-part tariff, a consumer’s water bill is based on the sum of two calculations. The single type of calculation used in a single-part tariff can be one of two types: a fixed charge or a water use (volumetric) charge; volumetric charges can be handled in several different ways. Figure 1 illustrates how the price of water to the consumer changes as the quantity of water used increases for some of these tariff structures. Figure 2 shows how the customer’s monthly water bill varies as the quantity of water used increases for selected tariff structures.

1.06.5.1 Single-Part Tariffs 1.06.5.1.1 Fixed charges In the absence of metering, fixed charges are the only possible tariff structure. With a fixed charge, the consumer’s monthly water bill is the same regardless of the volume used. In many countries, renters in multi-story apartment buildings have unmetered connections to their units and thus effectively pay

Table 8

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Basic types of water tariff structures

Single-part tariffs A. Fixed charge: monthly water bill is independent of the volume consumed B. Water use charge a. uniform volumetric tariff b. block tariff: unit charge is constant over a specified range of water use and then shifts as use increases (i) increasing block (ii) decreasing block c. increasing linear tariff: unit charge increases linearly as water use increases Two-part tariffs Fixed charge þ water use charge

a fixed charge for water (perhaps incorporated into their rent). Fixed charges are still quite widely used in industrialized countries, such as Canada, Norway, and the United Kingdom (and until recently in New York City), where water has historically been abundant and hence metering is not widespread. The fixed charge itself can vary across households or consumer classes depending on characteristics of the consumer. For example, historically a common way to charge differential fixed charges was to set higher fixed charges on more valuable residential properties, sometimes on the assumption that people living in higher-value dwellings tend to use more water and/or have a greater ability to pay for the water they use. It was also common to assign businesses, a different fixed charge than households, on the assumption that firms use more water than households, and notions of fairness (e.g., that firms have a greater ability to pay for water than households). Another common approach is to charge different monthly fees depending on the diameter of the pipe used by the customer to connect to the distribution system: single-family domestic connections generally require a smaller bore than connections for larger concerns (e.g., businesses, hospitals, and apartments). From the perspective of economic efficiency, the problem with a fixed-charge system is that consumers have no incentive to economize on water use, as using more water will not increase their water bill. If the short-run marginal cost of supply is very low due to excess capacity in the system, this may not be a big problem. However, from a cost recovery perspective, a fixed-charge system creates a potentially large problem for the utility (or operator) if some households still lack individual connections: customers who do have a connection can supply water to other users (e.g., unconnected households and vendors) without incurring an increase in the household water bill. Moreover, because the fixed charge offers no incentive to economize on the use of water, a fixed charge that provided sufficient revenues at one point in time will become increasingly inadequate as the economy and incomes grow and water use increases. W&S service providers will be reluctant to expand coverage because more customers may mean more financial losses. Fixed-charge tariffs are thus especially prone to locking communities into low-level equilibrium traps of few customers, low revenues, and poor service (Whittington et al., 1990).

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Pricing Water and Sanitation Services 1

Price ($/m3)

0.75

0.5

0.25

0 0

5

10

15

20

25

30

3)

Quantity (m Decreasing block

Increasing block

Increasing linear

Uniform

Figure 1 Price of water vs. the quantity of water used for selected tariff structures.

30

Monthly bill ($)

25 20 15 10 5 0 0

5

10

15

20

25

30

3)

Quantity (m Decreasing block

Increasing block

Increasing linear

Uniform

Figure 2 Monthly water bill vs. the quantity of water used for selected tariff structures.

1.06.5.1.2 Volumetric charges

1.06.5.1.3 Uniform volumetric charge

The second way to structure a single-part tariff is to base consumers’ water bills on the amount of water they use. In mathematical terms, the monthly water bill is thus a function of the quantity of water a consumer uses. The precise formula used for the calculation of the water bill can differ. There are three main options: (1) a uniform volumetric charge; (2) a block tariff where the unit charge is specified over a range of water use for a specific consumer, and then shifts as use increases; and (3) an increasing linear tariff whereby the unit charge increases linearly as water use increases. All volumetric charges require that the consumer has a metered connection and that this meter works reliably and is read on a periodic basis.

With a uniform volumetric charge, the household’s water bill is simply the quantity used (e.g., cubic meters) times the price per unit of water (e.g., US$ per cubic meter). This is the most common type of volumetric charge among water utilities in the United States, Australia, and a number of European countries and is also very common for industrial and commercial users throughout the world. A uniform volumetric charge has the advantage that it is easy for the consumer to understand, in part because this is how most other commodities are priced. From an economic efficiency point of view, it can be used to send a clear, unambiguous signal about the short-run marginal cost of using water.

Pricing Water and Sanitation Services 1.06.5.1.4 Block tariffs Block tariffs come in two main varieties: increasing and decreasing. They create a stepwise price structure as illustrated in Figure 1. With an increasing block tariff (IBT), consumers incur a low volumetric per-unit charge (price) up to a specified quantity (or block); for any additional water consumed, they pay a higher price up to the limit for a second block, even higher for the third, and so on. IBTs are widely used in arid areas such as Spain and parts of the Middle East, where water resources have historically been scarce. The use of IBTs is also widespread in many developing countries in Latin America and Asia. With a decreasing block tariff (DBT), on the other hand, consumers face a high volumetric charge up to the specified quantity in the first block, pay less per unit for additional water up to the limit for second block, then less still for the third, and so on. Thus, for both an IBT and a DBT structure, the water bill is calculated in the following manner: Let Q* ¼ amount of water sold to a specific consumer; Q1 ¼ maximum amount of water that can be sold in the first block at price P1; Q2 ¼ maximum amount of water that can be sold to a consumer in the second block at P2; Q3 ¼ maximum amount of water that can be sold to a consumer in the second block at P3. If Q*oQ1, then the consumer’s water bill ¼ (Q*) P1. If Q1oQ*oQ2, then the consumer’s water bill ¼ P1Q1 þ (Q* – Q1)P2. If Q1 þ Q2oQ*oQ3, then the consumer’s water bill ¼ P1Q1 þ P2Q2 þ (Q* – (Q1 þ Q2))P3. And so on for however many blocks there are in the tariff structure. The rationale commonly given for an IBT structure is that, in theory, it can achieve three objectives simultaneously. Proponents of an IBT argue that it promotes affordability by providing the poor with affordable access to a subsistence block of water (the lifeline rate). It can achieve efficiency by confronting consumers in the highest price block with the marginal cost of using water. It can raise sufficient revenues to recover costs. (Note: this argument assumes that the marginal cost of water is in fact higher than the first block price. But if a large expansion project has been recently completed, the short-run marginal cost of water may be very low.) The IBT structure has become so widely used in both industrialized and developing countries that many professionals working in the water sector assume that it must always be the most appropriate tariff structure. This is not the case. In practice, IBTs often fail to meet any of the three objectives mentioned above, in part because they tend to be poorly designed. An IBT may provide more expensive water to poorer households than to richer households, because in many cities the poor share connections, and in such cases the resulting higher volumetric use in turn results in higher prices for most of the water that those households consume (see, e.g., Whittington, 1992; Boland and Whittington, 2000; and Komives et al., 2005). Many IBTs also fail to achieve cost recovery and economic efficiency objectives, usually because the upper consumption blocks are not priced at sufficiently high levels and/or because the first subsidized consumption block is so large that almost all residential consumers never consume beyond that level. The DBT structure was designed to reflect the fact that when raw water supplies are abundant, large industrial

87

customers often impose lower average costs because they enable the utility to capture economies of scale in water-source development, transmission, and treatment. Also, large industrial users typically take their supplies from the larger trunk mains and thus do not require the expansion of neighborhood distribution networks. Although it is still used in some communities in the United States and Canada, the DBT has gradually fallen out of favor, in part because short-run marginal costs, properly defined, are now relatively high in some parts of the world, and there is thus increased interest in promoting water conservation by the largest customers. The DBT structure is also often politically unattractive because it results in high-volume users paying lower than average water prices.

1.06.5.1.5 Increasing linear tariff The increasing linear tariff structure is rarely used. It is of interest largely because it illustrates that there are many ways that water bills can be related to the quantity of water used. In this tariff structure, the price that a consumer pays per unit increases continuously (rather than in block increments) as the quantity of water used increases. (In other words, the water bill ¼ (Q*)P*, where Q* ¼ amount of water sold to a specific consumer; and P* ¼ (a1 þ a2)Q* and a1 and a2 are positive constants.) This tariff structure sends the consumer a powerful signal that increased water use is costly. Not only each additional unit of water used is sold at a higher price, but all the preceding units are sold at the last (high) price. A related but different tariff structure would require that only the last unit used would be sold at the highest price; other units would be sold at the price associated with the lower quantity. It is important to recognize, however, that an increasing linear tariff cannot send the proper economic signal to a consumer about the short-run marginal cost of additional water use. This is because the utility’s short-run marginal cost of providing water does not change appreciably as the water use of an individual household changes. An increasing linear tariff would thus be especially inappropriate if applied to large-volume industrial or commercial water users because it could drive the price they confront for increased water use far beyond the short-run marginal cost of supplying them additional water.

1.06.5.2 Two-Part Tariffs With a two-part tariff, the consumer’s water bill is based on the sum of two calculations: (1) a fixed charge and (2) a charge related to the amount of water used. There are many variations in the way these two components can be put together. The fixed charge can be either positive (a flat fee) or negative (a rebate). The water use charge can be based on any of the volumetric tariff structures described above (a uniform volumetric tariff, an increasing or decreasing block tariff, or an increasing linear tariff). In many cases, the fixed charge is kept uniform across customers and relatively low in value, and is used simply as a device for recovering the fixed administrative costs associated with meter reading and billing that are unrelated to the level of water consumption.

88

Pricing Water and Sanitation Services

1.06.5.3 Seasonal and Zonal Water Pricing In some circumstances the short-run marginal costs of supplying water to customers may vary by season. For example, a community may have relatively plentiful water supplies in the rainy season, but much more limited supplies in the dry season; water storage (reservoir capacity) will also be a factor. In such cases, it makes economic sense for water tariffs to reflect the varying circumstances. By charging higher rates in the dry season and lower rates in the wet season, water tariffs can be used to signal to customers that the water supply is not constant across the seasons, and that the costs of maintaining and distributing the water supply may vary as well. The higher dry season rate also serves as a reminder that each user’s consumption of water reduces the amount available for others. Chile is one of the few developing countries that currently uses seasonal water tariffs. Similarly, it may cost the water utility more to deliver water to outlying communities due to, for example, increased pumping costs to reach higher elevations or more distant settlements. A zonal water-pricing structure charges users who live in such areas more for their water because it costs the utility more to serve them. Zonal prices can be used as an economic signal to users that living in such areas involves substantially higher water supply costs and that such information should be factored into customers’ locational and water-use decisions. However, this practice is comparatively rare, in part because it requires the water supplier to collect detailed geographically referenced accounting information. This type of special tariff is only appropriate if the costs of serving the specially zoned areas are significantly higher than for the rest of the community. In fact, costs vary among all users, and a practical tariff always reflects averaged costs to some degree.

1.06.6 Achieving Economic Efficiency and Recovering Capital Costs: Fundamentals of Dynamic Marginal Cost Pricing in the W&S Sector The high costs of the capital investments necessary to build modern network W&S systems make the two-part tariffs described in the previous section especially attractive. They offer service providers a means simultaneously to achieve economic efficiency and cost recovery objectives and also to simplify the design of subsidies to aid poor households. Economic appraisal of W&S investments requires that stakeholders first determine the optimal price to charge for services, if the services are provided, and then determine whether the benefits are greater than the costs if the optimal price is charged. For large capital projects with no constraints on raw water supply, the volumetric charge (one component of a two-part tariff) may be very low in some circumstances because short-run marginal costs can be very low. The economic logic for setting price equal to the short-run marginal cost is straightforward (see, e.g., Layard and Walters, 1978, pp. 171–176). Consider a community with an inverse demand curve for W&S services p ¼ b1  b2x, where p ¼ price of the services, x is the quantity of W&S services that can

be supplied per time period, and b1 and b2 are positive coefficients. Let C equal the fixed costs per period of the W&S system, which is by definition assumed not a function of x. The investment is able to provide an amount of water Qc per period, where b1/b2 is less than Qc. Net benefits are maximized when the optimal quantity of W&S services x* is provided:

Total benefits  costs ¼

Z

x

ðb1  b2 xÞdx  C 0

¼ b1 x  12b2 x2 dðB  CÞ=dx ¼ b1  b2 x ¼ 0 x ¼ b1 =b2 Solving for the price that will achieve this optimal quantity, we see that in this simple example, in which there are no variable costs of water production and delivery and no opportunity costs of the raw water supply, the volumetric price should be set equal to zero (the short-run marginal cost):

p ¼ b1  b2 x p ¼ b1  b2 ðb1 =b2 Þ ¼ 0 If the price is set equal to zero to ensure customers receive the optimal quantity x*, the benefits of the project exceed the costs if

Z

Total benefits4 Costs b1 =b2

ðb1  b2 xÞdx4 C 0 1 2 2b1 =b2 4

C

Such a price will result in large financial deficits unless a fixed charge for capital recovery and other fixed costs is also imposed (the other component of a two-part tariff). The principles that a W&S service provider should follow to determine the volumetric and fixed-charge components of a two-part tariff in different circumstances have not been well understood in the water resources community. The key point is that short-run marginal costs change depending on the regional water resources situation, and both the volumetric and the fixed-charge components of the two-part tariff must change in response to changes in short-run marginal costs. A simple example can illustrate this point. Consider a community without a modern W&S network system that is thinking about undertaking a new project to develop such infrastructure along with arranging a new source of raw water supply. The various stakeholders consider the benefits and costs of such an infrastructure improvement and decide that the project is desirable (benefits exceed the costs). Assume that capital for this project is not available from a higher level of government or a donor agency; the city instead borrows the necessary funds from a bond market, promising to repay the loan from new revenues available from the sale of W&S services. The citizens of the community agree to allocate the responsibility for repaying this loan among all who use the W&S services. Upon the advice of the engineering firm

Pricing Water and Sanitation Services

responsible for designing the project, the community decides to build excess capacity into their W&S system in order to accommodate future population and economic growth. The engineers’ argument is that it is cheap to build this excess capacity now due to economies of scale in the various project components. Figure 3 presents the situation after this first project is built. The community now has the capacity to supply Qc. What volumetric price should the service provider charge? As the short-run cost of supplying additional water to the existing population is now low (because the capital costs have already been incurred), the volumetric charge should be low. The economic logic is that customers should not be discouraged from using more water if such use does not impose increased or significant costs on the W&S suppliers or neighbors. If a customer derives a benefit from using more water and this use does not hurt anyone else, why not permit the additional water use? However, the loan must still be repaid, so this consumer must pay a fair share of the capital costs. How a community determines a customer’s fair share is essentially a political matter. This decision will not affect the economically efficient outcome unless it significantly affects the number of customers who decide to connect to the W&S system. If large numbers of customers decide to disconnect from the system after a project is completed and the new tariff structure is imposed, in most cases this indicates a failure of the planning process. The voices of these customers were not likely to have been heard when the decision was made to build the project. A numerical example will help to clarify this argument. Assume that this W&S project has an average cost of US$0.75 per cubic meter and a short-run marginal cost of US$0.25 per cubic meter. Suppose that if the typical household were charged US$0.25 per cubic meter, it would use 20 cubic meters per month. In this case, the volumetric charge would yield revenue of US$5.00 per month (20 cubic meters  US$0.25 per cubic meter). However, when the loan repayment is considered, the utility actually needs US$15.00 per month from this household (US$0.75 per cubic meter  20 cubic meters). This implies that the fixed charge should be set equal to US$10.00 per month so that the utility can recover its average costs.

89

In this situation, the volumetric price sends a signal that there is more water available for households at this low shortrun marginal cost if households want to use it. The demand curve in period 1 (D1) intersects the short-run marginal cost curve at a point far below capacity Qc (Figure 3). Water is relatively abundant, and households, encouraged by a low volumetric charge, use plenty of water. They pay a significant fixed charge in order to repay the capital that they collectively agreed to borrow. In a well-governed community, households would have been made fully aware of the magnitude of the volumetric and fixed charges that would be necessary when they decided (voted) to undertake the new W&S project. Assume that this two-part tariff structure stays in place and that over time the population and economy of the community grow. As shown in Figure 4, the demand for W&S services shifts out and to the right. W&S services actually become more valuable to customers, but there is no need for the service provider to increase either component of the two-part tariff until point A in Figure 4 is reached, because the loan is being repaid and revenues are sufficient to pay the average costs of service. Customers thus enjoy an increasing consumer surplus on their W&S purchases. The citizens of this community are in effect reaping the benefits of their wise decision to invest in the new W&S project, and to include excess capacity into the project design. However, the water resources professionals can see that this excess capacity is being used up as growth continues, and the day will come when the community reaches the limits of its existing water situation (Qc). They make this known to the public. The community must then decide what to do before point A is reached, because it takes time to develop a new project. Essentially, it can either make do with the amount of water that it has (Qc) or build another water project. Suppose that there is another raw water source available to the community, but a project to develop this second source is more expensive than the one included in the first investment. Assume that this second project would result in a system-wide average cost of US$1.00 per cubic meter. Assume that the short-run marginal costs of the combined system (after this second project is built) would increase as well, to US$0.50 per cubic meter. Suppose that the community decides that the new, second project is too expensive (the benefits are less than the costs). The citizens vote against a bond referendum to raise money to

$0.50/m3 $0.50/m3

$0.25/m3

A

$0.25/m3 D1 D1

D2

Qc Qc Figure 3 First period: first project is completed, excess capacity.

Figure 4 Second period: demand grows as population and economic growth proceed; system capacity is reached at point A.

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undertake the new investment. Instead, they will try to make do with the water supply they already have. In this case the magnitude of the components of the two-part tariff must change, because the short-run marginal cost of using water changes. Now the volumetric charge must be used to ration the available water supplies, as shown in Figure 5. As population and economic growth proceed, the demand curve for water continues to shift up and to the right, but in this case the total quantity of water Qc available to the community is already being used by existing consumers (Figure 6). The shortrun marginal costs must now reflect the opportunity cost associated with taking water away from some customers: if one customer increases water use, another must decrease water use. The volumetric price of water thus keeps increasing to reflect the rising opportunity cost foregone (scarcity rent) and to ensure that the available water supply is accessible to users who need it most. Suppose that the community allows this process to go on, demand keeps growing, and the volumetric price needs to

$1.00/m3 B $0.50/m3 D3

$0.25/m3

A

Qc Figure 5 Third period: water from first source must be rationed.

$0.50/m3

D4

D5

D6

$0.25/m3

Qc Figure 6 Fourth period: demand continues to grow after completion of the second water project.

increase from US$0.25 to US$1.00 per cubic meter in order to ration the available supply. Assume that if the short-run marginal cost is US$1.00 per cubic meter and the service provider charges this price, average household consumption falls from 20 cubic meters per month to 12 cubic meters per month. Households economize on their use of water because the volumetric price has quadrupled. In effect, by cutting back on water use existing customers are leaving water available for new customers and new and expanded economic activities. This should not come as a surprise to existing consumers, because they themselves voted down the bond referendum that would have provided the finance for the second water project. Assuming that average costs do not change, the W&S service provider needs US$9.00 per month in revenue from the typical household customer (US$0.75 12 cubic meters). However, the volumetric charge yields US$12.00 in revenues (US$1.00 12 cubic meters). Most people would consider it unfair for the provider to reap windfall profits from the increase in the volumetric part of the tariff. The provider does not need the increased revenue to repay the loan or to pay its financial costs of operation. The purpose of the higher volumetric price is not financial, but rather to ration water use economically. The two-part tariff can be used to resolve this fairness problem associated with water rationing. The fixed charge should be reduced as the volumetric charge increases. Instead of a positive fixed charge of US$10.00, for example, a negative charge (rebate) of US$3.00 per month will result in a typical household water bill of US$9.00 per month, precisely the amount the provider needs to cover its costs. In this example, the fixed charge is negative, but this need not be the case. If the scarcity rent is small, the volumetric charge may not be large enough to recover the service provider’s costs, a positive fixed charge may still be needed. Now suppose that the rationing of the water available from this first project becomes an increasing burden on the citizens of the community, and they finally decide that it is worthwhile to build the second water project. This new project was projected to be more expensive than the first project, but nevertheless they vote to approve a bond referendum to finance it, because now they are paying a high volumetric price for water, US$1.00 per cubic meter, due to the high scarcity rent, and an increased supply will bring greater benefits than costs. Again, the community decides to build in excess reservoir capacity, to support further population and economic growth. After the second project is finished, what should the tariff be? The principles are the same as before: the volumetric component of the two-part tariff should be set equal to the short-run marginal cost, which has now risen from US$0.25 to US$0.50 per cubic meter. When the new water project opens, the citizens in this community are relieved that the constraint on their water use has been relaxed, and the volumetric price falls from US$1.00 to US$0.50 per cubic meter. Assume that in response to this decline in volumetric price, the typical household increases its water use to 16 cubic meters per month. Note that this volumetric price is less than in the previous period, when the price was being used to ration supplies, but more than in the first period, when the

Pricing Water and Sanitation Services

community was smaller and the first water project provided cheap and abundant supplies. If the typical household now uses 16 cubic meters and the volumetric charge is US$0.50 per cubic meter, the volumetric component of the two-part tariff yields US$8.00 per month. However, this is not enough for the W&S provider to recover its average costs, which now include loan payments on the second project. The total average cost of providing services is a weighted average of the first and second projects. Recall that this is assumed to be US$1.00 per cubic meter. In this case, the provider needs US$16.00 per month from the typical household (16 cubic meters  US$1.00 per cubic meter). The fixedcharge component of the two-part tariff must then be set at a positive US$8.00 per month. This example illustrates how a two-part tariff can be used to send the correct signal to customers about the economic value of water and at the same time address the financial needs of the W&S provider. The key point is that the volumetric charge should be continually adjusted to reflect the real short-run marginal cost of using water (including any opportunity costs associated with foregone uses), and the fixed-cost component should be adjusted to meet the financial needs of the utility. It is the community’s collective decision to agree (or not) to share the capital costs of the project that ensures that the benefits of the project exceed the costs and that the allocation of costs is considered fair by most parties. Note that regulatory authorities will have an important role to play in the establishment of an optimal two-part tariff. Particularly in times of water scarcity, when a high volumetric price and possibly a negative fixed charge is warranted, a regulatory body needs to ensure that the public understands the rationale for the pricing policy adopted. Unregulated private W&S service providers cannot be expected to reduce their fixed charge as the volumetric charge increases. The major objection to using a two-part tariff in this way is the possible instability in the volumetric price for services (in the example above, the volumetric price starts low, then quadruples, and then falls again). Some water-resource professionals and utility managers feel that changing volumetric prices will confuse customers and prevent them from engaging in careful long-range planning. From this perspective, price stability is a major objective of tariff design. Households and businesses are, however, able to deal with changing prices in the telecommunications and energy sectors, so there is reason to believe that these fears are exaggerated. Two-part tariffs are widely used in telecommunications pricing, although a negative rebate is not common because short-run marginal costs have continued to fall. (Actually, a form of negative rebate does occur with some mobile phone plans that provide expensive smart phones (e.g., Apple’s iPhone) to customers with long-term contracts at prices far below the cost that the mobile phone manufacturer sells the phone to the service provider.) Also, note that in some locations the period during which the volumetric price must be used to ration water use may be quite long. As cities need to go farther and farther afield in search of new supplies, managing water use with high prices may be increasingly attractive compared to incurring the rising capital costs of new projects. If the volumetric price of water cannot change in response to changing water resources’

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circumstances, it will be increasingly difficult to develop rational W&S pricing policies. But what about poorer members of the community? How can they be provided with improved W&S services when such a two-part tariff is used?

1.06.7 Subsidizing Capital Costs: Reaching the Poor Any discussion of W&S subsidies should begin with the question ‘‘Why do many people (both those working in the water supply sector and others elsewhere) assume that it is a good idea to deliver subsidies to the poor by reducing the water bills of households with private connections?’’ What is it about a piped water distribution network that makes it a good candidate for the delivery of subsidies to the poor? It does not follow that because water itself is a basic need, a piped water distribution system provides an efficient and an effective way to deliver subsidies to the poor. After all, people also have basic needs for food, health services, and housing. The relevant question is not ‘‘How can piped water services be subsidized most effectively?’’ but ‘‘Which subsidy mechanisms reach the poor most efficiently and effectively?’’ It is also important to ask how households themselves view the importance of the good or service to be subsidized. There is strong evidence that households indeed want improved W&S services as their incomes increase; this correlation between W&S coverage and household income suggests that these services are normal goods. As economic growth occurs in developing countries, more and more people obtain improved infrastructure services. Progress is being made particularly in China and India. Figure 7 shows the percentage of households at different income levels that have four infrastructure services (piped water, sewer, electricity, and telephone); the data come from interviews with more than 55 000 households in 15 developing countries (Komives et al., 2003). What is noteworthy about these households is that at all income levels, more people have electricity than have piped water or sewer. Very few of the poorest households have piped water or sewer, yet almost a third of those households have electric service. As monthly household income increases from very low levels to US$300 per month, coverage of all of these infrastructure services increases rapidly; above US$300, coverage increases at a slower rate. The data in Figure 7 should be interpreted carefully. It could be that more households have electricity because W&S networks were not, but electricity was, available in their neighborhoods, or because electric service was less expensive than W&S service. But in fact, monthly household bills for electricity are almost always higher than for W&S service; thus, a comparatively lower cost of service does not explain the pattern seen here, where many households have obtained electricity even when they do not have piped water. Figure 8 shows the percentage of households with different infrastructure services at different income levels in Kathmandu, Nepal. All of these households had the option to connect to all three network infrastructure services: electricity, water, and sewer. The majority of the very poor chose electricity, but not water and sewer. At higher income levels the percentage of households with W&S services is also higher, but

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the percentage of households with electricity is always higher still. The important point to recognize from these examples is that although water itself is a necessity, this does not necessarily mean that people prefer piped water service to electric service. Indeed, because water is a necessity, households must already have some water source. The question is thus how much an improved source is worth to them. This will depend on many factors, but probably the most important is how poor the household’s existing water service actually is. W&S planners often present the need for improved services as a moral imperative or a basic human right, but given the

choice, many households in developing countries would appear to want electricity before an in-house piped water or sewer connection. In fact, it is unusual for a household in a developing country to have a piped water connection and not have electricity. Figure 9 shows how the prevalence of different infrastructure bundles changes as household income increases. Almost no one, at any income level, has only a piped water service. However, many people do have electricity and not water. Many households in fact have no infrastructure services at all, although that percentage declines rapidly as household income increases. These data suggest that although most households would certainly like improved W&S services,

% hhs with electricity % hhs with sewer connection

% hhs with in-house water tap % hhs with telephone

100

% hhs with service

80

60

40

20

0 200

0

400

600

800

1000

1200

Median monthly household income in 1998 (US$) Figure 7 Infrastructure coverage vs. household income. From Komives K, Whittington D, and Wu X (2003) Infrastructure coverage and the poor: A global perspective. In: Brook P and Irwin T (eds.) Infrastructure for Poor People: Public Policy for Private Provision, ch. 3, pp. 77–124. London: World Bank and Public–Private Infrastructure Advisory Facility.

Electricity

% hhs with service (among hhs with access to 3 services)

100 90

In-house tap

80 70

Sewer

60 50 40 30 20 10 0

28

45

62

74

Median monthly household income in 1998 (US$) Figure 8 Infrastructure choices vs. household income: Kathmandu, Nepal.

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93

% of hhs with bundle in home

50 40 30 20 10

e hr e

ci t ri

Al lt

ty

t el ec d

p ta

p

an

ta In -h ou se

In -h ou se

ci

ty

an

an

d

d

to

to

ile

ile

t

on ly ta tri El ec

ci

ty

In -h ou se

t ri El ec

p

on ly

on ly le t To i

No ne

0

Figure 9 Household infrastructure bundles vs. household income: Asia vs. rest of the world. From Whittington D and Komives K (2002) The challenge of demand assessment in pro-poor infrastructure projects. Presentation at the PPIAF/ADB Conference on Infrastructure: Providing Solutions for the Poor – The Asian Perspective. Manila, November.

this is by no means their most important development priority. Given the choice, in some locations households would probably prefer to have any available subsidies directed to other sectors (e.g., roads, power generation, and education). But suppose that a city’s public health professionals and other development experts decide that W&S services are merit goods that must be subsidized. How best can this be done? In his memoirs (Yew, 2000), the former prime minister of Singapore, Lee Kwan Yew insightfully summarizes his philosophy: subsidize investment and savings, not consumption. He succinctly states the advantage of the two-part tariff with respect to making W&S services affordable to poor households. Subsidizing consumption by selling water at low volumetric prices without an accompanying fixed charge is a never-ending distortion, a signal that continually sends customers the wrong message about how expensive the fixed costs of W&S services really are. But once the capital costs are sunk, low volumetric prices may be appropriate to let consumers know the short-run marginal cost consequences of their decisions. This logic means that any available subsidies for piped W&S services should be directed to (1) lowering connection charges and (2) reducing the recurrent fixed charge component of the monthly bill. Capital subsidies are, of course, not without problems. In theory, if the political process can ensure that only economically sound public investments are undertaken, capital subsidies can both assist poor households and foster economic growth. But capital subsidies for infrastructure investments in general, and for W&S investments in particular, require disciplined public sector decision making. Such discipline is extremely difficult when subsidies come from outside the community that is to benefit from the investment. In most circumstances, a community would be foolish to decline a capital grant for an infrastructure project with an associated stream of positive benefits. It is the high initial costs that are

typically the hurdles to W&S improvements, and if someone else volunteers to pay these costs, why not let them? In practice, it has proved almost impossible for national governments or donor agencies to conduct rigorous economic appraisals of W&S projects. Whenever it appears that a particular project might not pass a cost–benefit test, water professionals appeal to intangible benefits to argue that the investment will in fact pass the test. This is particularly the case in the evaluation of rural W&S investments in developing countries, where neither donors nor national agencies attempt serious project appraisal of W&S projects. It is not hard to provide justifications for subsidies of social overhead capital such as water, transportation, and power infrastructure; the problem is removing such subsidies when they are no longer needed. As Hirschman pointed out, ‘‘The trouble with investment in social overhead capital (e.g., water and sanitation investments) y is that it is impervious to investment criteria. y As a result, social overhead capital is largely a matter of faith in the development potential of a country or region. y Such a situation implies at least the possibility of wasteful mistakes’’ (Hirschman, 1958, p. 84, emphasis added.) This is precisely what we have witnessed in the W&S sector, where white elephants and poorly performing projects have been a standard feature of the sector landscape (Therkildsen, 1988). When higher levels of government (or donor agencies) pay the capital costs of W&S projects, numerous opportunities for rent-seeking and corruption arise (Lovei and Whittington, 1993; Olson, 2000; Davis, 2004). If subsidizing the water bills of households connected to piped networks is a bad idea, what policies can instead or additionally be put into place to protect poor households from the rise in piped water bills that will be required for effective improvements and reforms? There are in fact a number of regulations or policy initiatives that can be coupled with the tariff structure to protect poor customers. The most

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obvious is simply to identify poor households and give them cash assistance to pay their water bills. This is essentially the approach now used in Chile. But even without such means testing, two sets of appropriate pro-poor policies are available.

1.06.7.1 Create a Well-Run System of Public Taps as a Safety Net for the Poor In every locale, W&S providers and regulatory bodies planning to install or expand a piped W&S system need to look carefully at any existing system of public taps. (Here, the term ‘public taps’ refers to a system of fountains in public areas outside of people’s residences where anyone can go to collect water, perhaps for a per-bucket charge or a fixed monthly fee. These public taps do not necessarily need to be run by a public sector utility; they could be efficiently built and managed by a private operator.) In many places, public taps will become obsolete if and when piped services become available: where the majority of households have piped water connections, households without private connections will work out efficient ways of obtaining water from their neighbors at relatively low cost (Whittington et al., 1998). This solution depends on improving the piped distribution system so that connected households do not have to worry about running out of water if they give or sell water to their neighbors. (Public taps will become relatively high-cost sources of supply compared to purchasing from neighbors, because most unconnected households will have to walk farther to collect water from public taps than to obtain it from neighbors, and because the fixed costs of an attendant at the public tap will be large relative to revenues if only low volumes of water are sold.) Nevertheless, public taps may still have an important role to play because they may serve as a water source of last resort for the very poor. In some cases, it is even possible to provide water free from public taps without substantially reducing the revenues of the water utility. This can occur when the availability of free water from public taps does not reduce the number of households desiring private connections for their exclusive use, and when only a small number of households cannot afford private connections. (This is, in fact, the situation in many industrialized countries today. Water is often available free from public fountains, but the vast majority of households still demand private connections in their residences (see, World Bank Water Demand Research Team, 1990).) One source of potential revenue for financing a subsidized system of public taps is the excess revenues that are available if the volumetric price of water from private connections is higher than average costs.

1.06.7.2 Preserve Options for the Poor Poor households are hurt most when they have few options for self-help and when others have restricted their choices. In such cases, it is common to find poor households being exploited. This is as true in the W&S sector as elsewhere. One important way to protect poor households is to preserve their choices so that local mafia or other rent-seeking actors cannot exploit them. There are three main things that can and should be done.

1. Ensure that poor households (and others) can have a private water connection when they want it. Pro-poor policies should not trap poor households into always accepting a low level of off-site water service. If a poor household always has the option of choosing a private connection, when they can afford it, there are limits to the degree they can be exploited by rent seekers. 2. Legalize water vending and sale of household water to neighbors. Vendors and neighbors with private connections create options for poor households: they promote competition in local water markets, limit the reach of spatial monopolies, and drive down water prices. The poor will benefit most from these lower prices. The system of public taps described above also adds to the choices available to poor households, fosters competition, and thus protects the poor from exploitation. 3. Do not give private operators exclusive rights to provide water within a service area. Contracts with private operators should not contain exclusivity clauses. These limit competition and typically end up restricting poor households’ options. Small-scale providers can often lower the cost of providing piped water to poor households; they should be permitted to operate within the contract areas of larger private operators.

1.06.8 Concluding Remarks Two-part tariffs have an important role to play in enabling water utilities simultaneously to achieve economic efficiency and cost recovery objectives. If a large-capacity expansion project has recently been completed, the short-run marginal cost of raw water supply may be very low. Economic efficiency requires that water should be priced at short-run marginal cost. If a two-part tariff is used, however, the necessary revenues can be raised via a fixed charge, without distorting the price signal contained in the volumetric charge. However, in periods of water scarcity (e.g., just before the construction of a water supply augmentation project), the situation is reversed. In this case, pricing at short-run marginal cost implies that the volumetric charge must include the opportunity cost to the user who does not receive water due to scarcity. This scarcity rent causes the volumetric charge to be relatively high in order to ration the available water supply among competing users. Such high volumetric charges may produce revenues in excess of financial costs. This can be corrected by employing a negative fixed charge (rebate), while the volumetric charge remains high enough to send the correct signal to customers from an economic efficiency perspective. Such dynamic tariff design will require that W&S service providers, regulatory bodies, and public officials provide much more information to customers on the rationale behind sound pricing policies. As Hanemann (2005) has observed, it is extremely difficult for publicly owned W&S utilities to receive permission from political regulatory authorities for even modest rate increases, even though such increases are routinely granted to other service providers such as cable television. As water resources management becomes increasingly complicated, the public must become better informed about the challenges for tariff design posed by the high capital costs

Pricing Water and Sanitation Services

of W&S services, the long lives of the projects, and the tradeoffs between competing objectives. This degree of public understanding is unlikely to happen without increasing involvement and participation of stakeholders in the water resources planning and investment process in general and tariff design, and in rate setting in particular.

References Bahl RA, Sinha C, Poulos D, et al. (2004) Costs-of-illness of typhoid fever in Indian urban slum community: Implications for vaccination policy. Journal of Health, Population, and Nutrition 22(3): 304--310. Baumann D and Boland J. (1998) The case for managing urban water. In: Baumann D, Boland J, and Hanemann W. (eds.) Urban Water Demand Management and Planning, ch. 1, pp. 1–28. New York, NY: McGraw-Hill. Boland J (1993) Pricing urban water: Principles and compromises. Water Resources Update 92: 7--10. Boland J and Whittington D (2000) The political economy of increasing block water tariffs in developing countries. In: Dinar A (ed.) The Political Economy of Water Pricing Reforms, pp. 215--236. Oxford: Oxford University Press. Davis J (2004) Corruption in public services delivery: Experience from South Asia’s water and sanitation sector. World Development 32(1): 53--71. Davis J, Kang A, Vincent J, and Whittington D (2001) How important is improved water infrastructure to microenterprises? Evidence from Uganda. World Development 29(10): 1753--1767. Engineering News-Record (2004) 253. 24 (December 12), 32–37. Esrey SA and Andersson I (2000) Ecological sanitation – a missing link to sustainable urban development. Paper presented at the International Symposium ‘‘Urban Agriculture and Horticulture: The Linkage with Urban Planning.’’ Doma¨ne Dahlen, Berlin, Germany, 7–9 July. Hanemann WM (2005) The economic conception of water. In: Peter P, Rogers M, Llamas R, and Martinez-Cortina L (eds.) Water Crisis: Myth or Reality, pp. 61--91. London: Taylor and Francis. Hirschman A (1958) The Strategy of Economic Development. New Haven, CT: Yale University Press. Komives K, Whittington D, and Wu X (2003) Infrastructure coverage and the poor: A global perspective. In: Brook P and Irwin T (eds.) Infrastructure for Poor People: Public Policy for Private Provision, ch. 3, pp. 77–124. London: World Bank and Public–Private Infrastructure Advisory Facility. Komives KV, Halpern FJ, and Wodon Q (2005) Water, Electricity, and the Poor: Who Benefits from Utility Subsidies? Directions in Development. Washington, DC: World Bank. Layard PRG and Walters AA (1978) Microeconomic Theory. New York, NY: McGrawHill. Lovei L and Whittington D (1993) Rent-seeking in the water supply sector: A case study of Jakarta, Indonesia. Water Resources Research 29(7): 1965--1974. Middleton R, Saunders R, and Warford J (1978) The costs and benefits of water metering. Journal of the Institution of Water Engineers and Scientists 32: 11--122. Nauges C and Whittington D (2009) Estimation of water demand in developing countries: An overview. The World Bank Research Observer (forthcoming). Olson M (2000) Power and Prosperity: Outgrowing Communist And Capitalist Dictatorships. New York, NY: Basic Books.

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Pattanayak S, Yang J, Whittington D, and Kumar B (2005) Coping with unreliable public water supplies: Averting expenditures by households in Kathmandu, Nepal. Water Resources Research 41(2): W02012. Poulos C and Whittington D (2000) Individuals’ rates of time preference in developing countries: Results of a multi-country study. Environmental Science and Technology 43(8): 1445--1455. Sara J, Gross A, and Berg C (1996) Rural Water Supply and Sanitation in Bolivia: From Pilot to National Program. UNDP–World Bank Water and Sanitation Program. Therkildsen O (1988) Watering White Elephants: Lessons from Donor-Funded Planning and Implementation of Rural Water Supplies in Tanzania. Uppsala: Scandinavian Institute of African Studies. United Nations Human Settlement Programme (UN-HABITAT) (2003) Water and Sanitation in the World’s Cities: Local Action for Global Goals. London: Earthscan. United Nations Millennium Project Task Force for Water and Sanitation (2004) What will it take? Water, sanitation, and the millennium development goals. Abridged Draft Final Report. November. New York, NY: SIWI. Van Wijk-Sijbesma C (1989) What Price Water? User Participation in Paying for Community-Based Water Supply. Occasional Paper Series, No. 10. The Hague: International Reference Centre for Community Water Supply and Sanitation. Warford J (1994) A marginal opportunity approach to municipal water pricing. EEPSEA Special Paper. Singapore. White G, Bradley D, and White A (1972) In: Drawers of Water: Domestic Water Use in East Africa. chs. 4 and 5. Chicago: University of Chicago Press. Whittington D (1992) Possible adverse effects of increasing block water tariff in developing countries. Economic Development and Cultural Change 41: 75--87. Whittington D (2003) Municipal water pricing and tariff design: A reform agenda for South Asia. Water Policy 5: 61--76. Whittington D (2006) Reflections on the goal of universal access in the water and sanitation sector: Lessons from Ghana, Senegal, and Nepal. In: Liberalisation and Universal Access to Basic Services: Telecommunications, Water, Sanitation, Financial Services, and Electricity, chap. 5, pp. 135–148. OECD, The World Bank. Paris: OECD Publishing. Whittington D, Boland J, and Foster V (2002) Water Tariffs and Subsidies in South Asia: Understanding the Basics. New Delhi: Water and Sanitation Program. Whittington D, Davis J, Komives K, et al. (2009a) How well is the demand-driven, community management model for rural water supply systems doing? Evidence from Bolivia, Peru, and Ghana. Water Policy 11(6): 696--718. Whittington D, Davis J, and McClelland E (1998) Implementing a demand-driven approach to community water supply planning: A case study of Lugazi, Uganda. Water International 23: 134--145. Whittington D, Hanemann WM, Sadoff C, and Jeuland M (2008) The challenge of improving water and sanitation services in less developed counties. Foundations and Trends in Microeconomics 4(6–7): 469–609. Whittington D and Komives K (2002) The challenge of demand assessment in propoor infrastructure projects. Presentation at the PPIAF/ADB Conference on Infrastructure: Providing Solutions for the Poor – The Asian Perspective. Manila, November. Whittington D, Okorafor A, Okore A, and McPhail A (1990) Strategy for cost recovery in the rural water sector: A case study of Nsukka district, Anambra state, Nigeria. Water Resources Research 26(9): 1899--1913. World Bank Water Demand Research Team (1990) The demand for water in rural areas: Determinants and policy implications. The World Bank Research Observer 8(1): 47--70. Yew L (2000) From Third World to First – the Singapore Story: 1965–2000. Singapore: Times Publishing Group.

1.07 Groundwater Management E Lopez-Gunn, Complutense University, Madrid, Spain MR Llamas, Complutense University, Madrid, Spain A Garrido, Polytechnic University of Madrid, Madrid, Spain D Sanz, University of Barcelona, Barcelona, Spain & 2011 Elsevier B.V. All rights reserved.

1.07.1 1.07.2 1.07.2.1 1.07.2.2 1.07.2.3 1.07.2.4 1.07.2.5 1.07.2.5.1 1.07.2.5.2 1.07.2.6 1.07.3 1.07.3.1 1.07.3.2 1.07.3.3 1.07.3.4 1.07.3.5 1.07.3.6 1.07.3.7 1.07.4 1.07.4.1 1.07.4.1.1 1.07.4.2 1.07.4.3 1.07.5 1.07.5.1 1.07.5.1.1 1.07.5.1.2 1.07.5.2 1.07.5.2.1 1.07.5.2.2 1.07.6 1.07.6.1 1.07.6.1.1 1.07.6.1.2 1.07.6.1.3 1.07.6.2 1.07.6.2.1 1.07.6.2.2 1.07.6.2.3 1.07.7 References

Introduction The Global Silent Revolution of Intensive Groundwater Use Introduction The Role of Groundwater in the Global Water Cycle Location of the Main Aquifers Groundwater Uses: Past and Present The Pros and Cons of the Intensive Use of Groundwater The complex meaning of sustainability in groundwater use The ethics of pumping nonrenewable groundwater (groundwater mining) The Social Sustainability of Groundwater Management The Economics of Groundwater Use Groundwater Costs of Abstraction and Groundwater Tariffs Productivity of Groundwater Use Perverse Subsidies in Water Policy Groundwater: From Open Access to Common Pool? Optimal Groundwater Pricing Departures from Optimality: Second-Best Solutions Internalizing the Value of Environmental Services Provided by Groundwater Regulatory Frameworks for Groundwater Multilevel Governance Diversity in Groundwater Regulatory Regimes The controversy over private, public, or community groundwater rights Implementation and Enforcement of Groundwater Legislation Multilevel Regulatory Frameworks Institutional Aspects of Groundwater Management Groundwater Institutions: Mapping Groundwater Institutional Design Boundary definition The role of groundwater-user associations An Institutional Audit of Groundwater Institutions Higher-level authorities: Supporting, legitimizing, and leading Transparency and participatory groundwater management The Complex Concept of Groundwater Sustainability and Future Management Issues Groundwater Management Externalities Degradation of groundwater quality Susceptibility to subsidence Interference with surface water and ecological impacts Groundwater: Future Risks and Opportunities for Management Groundwater and climate change Future management issues Groundwater: Issues of fit and political windows of opportunity Conclusion

1.07.1 Introduction The last half century has witnessed a spectacular development in groundwater use. Its use has increased from some 100 Mm3 to almost 1000 Mm3 in the period 1950–2000 (Shah, et al., 2007). In this chapter, water volumes are referred to in million cubic meter (106 or Mm3) and billion cubic meter (109, km3 or bcm).

97 98 98 99 99 99 99 100 100 103 103 104 104 105 105 106 106 108 110 110 111 112 113 114 114 114 116 117 117 117 118 119 119 120 120 121 121 123 123 124 124

The table of conversions is as follows:

• • • •

Mm3 ¼106 m3 km3 ¼109 m3 bcm ¼ 109 m3 hm3 ¼106 m3.

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Groundwater Management

This groundwater development has been mainly used for irrigation in arid and semiarid regions. Nevertheless, the use of groundwater for rural and urban water supply is very important; in some countries, like in Italy, it represents 90% of all of the urban water supply. Although the accuracy of available water-use data is still rather illusory, it seems that today the economic value of groundwater irrigation is even greater than the corresponding value of surface-water irrigation. The main factors that have driven this spectacular development in groundwater use include (1) the availability of modern and relatively cheap rigs to drill water wells; (2) the invention and ease of use of the turbine pump that allows abstracting significant volumes of groundwater from deep water wells; and (3) the consolidation of hydrogeology as a reliable science and technology that has dispelled the mystery of groundwater (Figure 1). This extraordinary development of groundwater use has been described as a silent revolution (Forne´s et al., 2005; Llamas and Martı´nez-Santos, 2005) because it was due to the efforts of millions of farmers, with scarce planning and control by conventional water authorities. These public government bodies have been occupied for more than 50 centuries with surface-water systems, beginning in the hydraulic civilizations located in the valleys of large rivers, such as the Nile, the Ganges, and the Yellow River. Therefore, it is not surprising that most high-level water decision makers suffer from hydro-schizophrenia. This disease was described for the first time in 1973 by the American hydrologist Raymond Nace, as the mindset of those that completely separate surface water and groundwater, and usually forget the latter (Llamas, 2004).

From the dug-well to the deep borehole.

This silent revolution has produced great benefits to humankind because it has contributed significantly both to reduce malnourishment in poor countries and to provide drinking water to the rural and urban poor. Moreover, groundwater irrigation is generally a driver for positive social changes. However, the current and frequent situation of inadequate planning and control over groundwater development has also triggered problems, which are mainly related to negative ecological impacts on aquatic ecosystems and groundwater-quality degradation. These problems have been frequently exaggerated by many surface-water experts who have created the pervasive hydromyth of groundwater fragility in order to foster the traditional policy of surface-water infrastructure. In summary, our main message is that it is crucial that high-level water decision makers seriously consider the real role that groundwater is playing and can play in current and future water policy. This role is going to increase even more if the predictions by the International Panel for Climate Change (IPCC) of an increase in temperature and a decrease in precipitation in most arid and semiarid regions become true (Bates et al., 2008). Our emphasis in this chapter is on groundwater management and not on groundwater hydrology. This is, first, because one of the volumes in this treatise deals with hydrology; second, because in our view, the main current problem is not lack of knowledge about aquifer location, characteristics, and functioning, but rather about better ways to manage the aquifers as a common-pool resource. We refer the interested reader to the work of Chevalking et al. (2008) for some useful citations of websites pertaining to the various aspects of groundwater management.

From the water wheel to the pump.

Figure 1 The silent groundwater revolution: changes in global water management.

From the water-witches to hydrogeology.

Groundwater Management

Consequently, after this introduction, Section 1.07.2 is devoted to recall the value of groundwater as a strategic resource. The main specific characteristics of groundwater that require a different management style than surface water is emphasized. Data provided by the United Nations Economic Scientific and Cultural Organization (UNESCO)–International Groundwater Resources Assessment Centre (IGRAC), the International Association of Hydrogeologists and the book by Margat (2008) describe the location of the main aquifers and groundwater uses in most countries. Developing (or semi-developing or emerging) arid and semiarid regions, such as India or regions of East Asia, have experienced a spectacular development in groundwater irrigation during recent years (Shah, 2005; Shah et al., 2007). Such large regions may present a wide variety of conditions: from subsistence livelihoods to market economies and from large alluvial aquifer systems, which may sustain long-term groundwater development, to hard-rock aquifers, where small communities may rely on scarce resources and pumping may prove to be costly. Some arid regions are endowed with good aquifers: these may correspond to countries, such as Saudi Arabia or Libya, where groundwater mining is commonplace (see Section 1.07.2.5.2). Reliance on nonrenewable resources, however, does not seem to render these economies unsustainable. Contrary to the perception of some environmental organizations, a good number of authors and the UNESCO World Commission on the Ethics of Science and Technology (COMEST) consider the use of nonrenewable groundwater resources to be acceptable under certain circumstances (Selborne, 2001; Delli Priscoli et al., 2004; Llamas, 2004). In arid and semiarid regions of industrialized countries, for example, the USA (California, Texas) and Spain, intensive groundwater withdrawals for irrigation are a well-established practice. Development is essentially market driven, as the cost of obtaining groundwater generally amounts to a very small fraction of the crop value. Some authors argue that the depletion of groundwater levels results in an increase of pumping costs, and may ultimately yield these intensive uses economically unsustainable. However, empirical evidence in some areas seems to show the opposite. Farmers are not deterred from pumping despite depths in excess of 400 m (Garrido et al., 2006). This is because switching to highervalue, water-efficient crops may offset the increase of pumping costs, provided that groundwater quality does not worsen (Llamas and Martı´nez-Santos, 2005; Forne´s et al., 2005). It can also be explained by the so-called Gisser–Sanchez effect (see Koundouri (2004) for a detailed explanation), that is, ‘‘the nomanagement (competitive) dynamic solution of groundwater exploitation is almost identical (in terms of derived social welfare) to the efficient management (optimal control) solution’’ (p. 706). This is a management paradox because the serious depletion of aquifers is a major risk to many freshwater ecosystems; yet, the social benefits from managing groundwater extraction are numerically insignificant. This also has significant implication for water managers because it severely constrains the effectiveness of policy options, since implementing reduced extractions is not socially, economically, or politically costless.

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1.07.2 The Global Silent Revolution of Intensive Groundwater Use 1.07.2.1 Introduction This section summarizes the main hydrologic characteristics of groundwater occurrence, availability, and past and present uses. A detailed description of these aspects can be found in Margat (2008), in the UNESCO–IGRAC website, and is also treated in one of the volumes in this treatise. Intensive use of groundwater is a recent phenomenon, less than half a century old in most places. This situation has occurred mainly in arid and semiarid countries, in some coastal zones, and close to a few mega-cities. This groundwater development has produced great socioeconomic benefits, mainly in developing countries. It has provided cheap drinking water that has helped improve public health. The new irrigated lands have contributed to eradication, or at least mitigation, of malnourishment among those living in poverty. Millions of modest farmers with scarce public or governmental planning, assessment, financing, and control have mainly carried out this intensive groundwater development. This intensive use has really been a kind of silent revolution. In most countries, the corresponding public water or irrigation agencies have been mainly devoted to designing, building, and operating large surface-water irrigation systems. The attitude of some water decision makers who strongly separate surface and groundwater projects, usually ignoring groundwater, was described as hydro-schizophrenia by the well-known American hydrologist, Raymond Nace, in the year 1973 (Llamas, 2004). This attitude has been commonplace in India, Mexico, Spain, and many other arid and semiarid regions worldwide. As a consequence, certain adverse effects have ensued in some places. For instance, in South Asia, the current situation concerning groundwater development has been frequently described as colossal anarchy (Shah et al., 2007). Most of the problems caused by this uncontrolled groundwater development could be avoided or mitigated if the corresponding government agencies had been more active in assessing and controlling groundwater use. On the other hand, surface-water officials have frequently exaggerated such problems. This has created a pervasive hydromyth on the fragility or weakness of groundwater as a reliable resource (Custodio, 2002; Lopez-Gunn and Llamas, 2008). Due to ignorance, vested interests, or, more frequently, because of the low credibility of the official warnings of watergoverning bodies about potential threats, most farmers are not reducing their intensive groundwater abstraction. On the other hand, there are practically no documented cases where intensive groundwater abstraction from medium- or largesized aquifers has caused serious social or economic problems similar to those caused by soil water logging and salinization, or by the people displaced or ousted by the construction of large dams.

1.07.2.2 The Role of Groundwater in the Global Water Cycle The inventory and the movement of water on planet Earth is well known and acceptably quantified, for at least a half century. According to Margat (2008), groundwater storage

100

Groundwater Management

globally, that is, the volume of groundwater stored in the geological formations defined as aquifers, is huge, about 107 km3. This is about 98–99% of all the liquid freshwater in the Earth, although it is only about 1% of the total volume on the hydrosphere (including oceans). The hydrological cycle indicates that water is continuously in motion, driven mainly by solar energy and gravity. This flow of water is very important. It is estimated that every year, precipitation is in the order of 100 000 km3, river flow about 40 000 km3, and groundwater recharge in the order of 10 000 km3. This means that the storage is 1000 times greater than the annual recharge. A large part of this groundwater recharge or runoff feeds the rivers and is included in river flow. However, these global numbers give only a preliminary idea and the specific situation in each region may be quite different. In Margat (2008) and in the UNESCO–IGRAC website, details can be found by country and continent. The main idea to keep in mind is that the volume of freshwater stored in the aquifers is usually huge in comparison with their yearly recharge (usually coming from precipitation) and with discharge (usually to rivers or wetlands). The annual recharge may change from practically zero in the most arid regions to more that 1000 mm yr1 in very humid areas.

1.07.2.3 Location of the Main Aquifers Most countries have made significant efforts to characterize the geological formations defined as aquifers, including their main parameters and functioning and relation with surfacewater bodies. These data are usually synthesized in different types of maps, known as hydro-geological maps. The UNESCO–IGRAC is a center that collects these data from all over the world and makes them available to the general public. The interested reader can find a good summary in the work by Margat (2008) of the situation in most countries (Figure 2) and a brief description of the main aquifer systems, with special emphasis on those that are transboundary (Figure 3). This means that they occupy areas in two or more countries. As an international transboundary resource, it is estimated that there are nearly 240 transboundary groundwater systems or aquifers (WHYMAP, Worldwide Hydrogeological Mapping and Assessment Programme; Lopez-Gunn, 2009). Margat (2008) classifies the aquifers into four groups according to their main geological characteristics: karstic, alluvial, hard rock, and volcanic. Significant attention is devoted to the aquifers in arid and semiarid regions because in these areas, the recharge of groundwater is small and its use may be relevant.

1.07.2.4 Groundwater Uses: Past and Present The use of groundwater coming from springs is as old as humanity. Margat (2008) mentions historical data of dug wells several millennia old. In Armenia, the first infiltration galleries to drain phreatic aquifers are recorded dating to 8 BC. This technology was soon extended to the whole Mediterranean region and to Asia. Hundreds of thousand kilometers of these galleries were constructed, some of them still in operation today. Nevertheless, the spectacular increase in the use of

groundwater has been mainly driven by the improvement in drilling technology and the invention of the turbine pump in the first-third of the twentieth century. Margat (2008) estimates that groundwater abstractions in 2004 were 800 km3 yr1; Shah et al. (2007), however, estimate that this amount is more than 1000 km3 yr1 and its use is on the increase. For the sake of comparison recent studies on the water footprint and virtual water trade (Aldaya et al., 2009), the total use of green and blue water is estimated as 7000 km3 yr1; of this, probably blue water accounts for about 3000 km3.

1.07.2.5 The Pros and Cons of the Intensive Use of Groundwater As mentioned in Section 1.07.1, a good number of authors emphasize the problems related to groundwater development. In this chapter, we intend to present an objective appraisal. Groundwater development produces great economic benefits due mainly to its general resilience to drought, and thus allowing supply to meet demand in a timely fashion. The fact that most groundwater development has been done by private persons, mostly modest farmers, with no, or a small public subsidy, is the best evidence. Some externalities of groundwater development have negative impacts. These externalities are described in Section 1.07.6.1. Nevertheless, in agreement with Llamas and Custodio (2003), in general, many of the negative externalities have been exaggerated and/ or could have been corrected or mitigated by good groundwater management.

1.07.2.5.1 The complex meaning of sustainability in groundwater use Whenever adverse effects of groundwater development begin to be felt, it is common to hear about ‘overexploitation,’ a term usually equated to pumping in excess of the recharge. While this practice is often dismissed as unsustainable, the concept of overexploitation is conceptually complex. This is the reason why a significant number of authors consider it simplistic and potentially misleading (Selborne, 2001; Delli Priscoli et al., 2004; Llamas, 2004). Probably, the most complete analysis is the one by Custodio (2002). As a consequence, more and more authors are changing to the expression intensive use of groundwater instead of using groundwater overexploitation. Intensive groundwater use denotes significant changes on natural aquifer dynamics (Llamas and Custodio, 2003). In contrast with aquifer overexploitation, intensive groundwater use does not convey a positive or negative connotation. It merely refers to a change in flow patterns, groundwater quality, or interrelations with surface-water bodies. It has been stated that the frequently encountered view – that the water policy of arid countries should be developed in relation to renewable water resources – is unrealistic and fallacious. Ethics of long-term water-resources development must be considered with ever-improving technology. It has been customary – as in the Spanish 1985 Water Law – to define overexploitation as the situation when groundwater withdrawal exceeds or is close to the natural recharge of an aquifer. The observation of a trend of continuous significant decline of the levels in water wells during several years is frequently considered as a clear indication of an unsustainable

Rio de Janeiro

Cairo

Medium groundwater recharge (15 − 150 mm/a) Low groundwater recharge (< 15 mm/a)

Medium groundwater recharge (15 − 150 mm/a)

Low groundwater recharge (< 15 mm/a)

2000

3000

4000

Dhaka

5000 km

Jakarta

Seoul

Shenyang

Sydney

Selected city

Continuous ice sheet

Large saltwater lake

Large freshwater lake

Major river

Surface water and Geography

© BGR Hannover / UNESCO Paris 2006

Tokyo Osaka

Manila

Hong Kong

Shanghai

Tianjin

Beijing

Madras Bangkok

Hyderabad

Calcutta

Delhi

Bangalore

Bombay

Lahore Karachi

Area with local and shallow aquifers

1000

Tehran

Figure 2 Groundwater resources of the world. From World-wide Hydrogeological Mapping and Assessment Programme (WHYMAP), special edition 2006.

High groundwater recharge (> 150 mm/a)

0

Istanbul

Moskva

Saint Petersburg

Kinshasa

Area with complex hydrogeological structure

Buenos Aires

Sao Paulo

Lagos

High groundwater recharge (> 150 mm/a)

Santiago

Lima

Bogota

New York

Paris

Major groundwater basin

Groundwater

Mexico City

Los Angeles

Chicago

London

Ruhr area

Figure 3 Map on transboundary aquifers of the world. From World-wide Hydrogeological Mapping and Assessment Programme (WHYMAP), special edition 2006.

Groundwater Resources of the World

Groundwater Management

situation. This is a simplistic approach that might be a long way from the real situation. It often corresponds to a transient state of the aquifer toward a new equilibrium (Custodio, 2002). Intensive groundwater use frequently depletes the water table. Depletions of the order of 0.5 m yr1 are frequent, although rates up to 5–10 m yr1 have been reported (Llamas and Custodio, 2003; Garrido et al., 2006). Farmers are seldom concerned with this issue, except in the case of shallow aquifers. The increase in pumping costs is usually a small problem in comparison with potential groundwater-quality degradation or equity issues such as the drying up of shallow wells or khanats (infiltration galleries), owned by the less-resourceful farmers and located in the area of influence of the deep wells (Wegerich, 2006). This may cause social-equity problems in regions where many farmers cannot afford to drill new wells, or the water authorities are not able to demand just compensation in terms of water or money to poor farmers. The opposite phenomenon (rise of the water table due to surface-water over-irrigation) is also a problem, for example, in Punjab, India, and in Pakistan, or in San Joaquin Valley in California. Raising the water table often results in significant social and economic troubles due to soil waterlogging and/or salinization. It is not easy to achieve a virtuous middle way. As Collin and Margat (1993) state: ‘‘we move rapidly from one extreme to the other, and the tempting solutions put forward by zealots calling for Malthusian under-exploitation of groundwater could prove just as damaging to the development of society as certain types of excessive pumping.’’ In a given aquifer, pumping rates for irrigation may prove to be sustainable from the hydrological viewpoint provided that storage and/or average recharge are large enough. However, water table drawdown may induce degradation of valuable groundwater-dependent ecosystems, such as wetlands, which may be considered unsustainable from the ecological point of view. Would a restraint from pumping be the most sustainable course of action? The answer to this question is difficult. If farmer livelihoods rely heavily on groundwater resources, a ruthless push toward wetland restoration may not be the most sensible solution to the problem. In that case, like in many real-life situations, the social and economic aspects of sustainability come into play, and may eventually offset environmental considerations. Llamas et al. (2007) provide a succinct overview of nine different aspects of groundwater sustainability: hydrological, ecological, economic, social, legal, institutional, inter- and intra-generational, and political. Throughout the text, a distinction is often made between developed and developing regions. This is because perceptions as to what is sustainable vary across geographical boundaries, and are often rooted in cultural, political aspects, and the socioeconomic situations. In this regard, the Hydrogeology Journal theme issue of March 2006 (Llamas et al., 2006) presents the socioeconomic analyses of a number of case studies from all over the world. Therefore, any study on economic sustainability of groundwater use should take into account the specific regional settings. In developing countries where easily accessible unconfined shallow aquifers exist, devices such as the treadle pump to access shallow water tables may constitute a catalyst

103

for irrigation development, while environmental concerns are generally subordinated to human development. This is the case in many small African villages.

1.07.2.5.2 The ethics of pumping nonrenewable groundwater (groundwater mining) Some arid regions have very small amounts of renewable water resources but huge amounts of fresh groundwater reserves, for example, the existing reserves under most of the Sahara desert. In such situations, groundwater mining may be a reasonable action if various conditions are met: (1) the amount of groundwater reserves can be estimated with acceptable accuracy; (2) the rate of reserve depletion can be guaranteed for a long period, for example, from 50 to 100 years; (3) the environmental impacts of such groundwater withdrawals are properly assessed and considered clearly less significant than the socioeconomic benefits from groundwater mining; and (4) solutions are envisaged for the time when the groundwater is fully depleted. Selborne (2001), former chairman of the COMEST, agrees with this approach. In Saudi Arabia, the main aquifers (within the first 300 m of depth) contain huge amounts – a minimum of 2000 km3 – of fresh fossil water that is 10 000–30 000 years old. It is considered that these fossil aquifers can supply useful water for a minimum period of 150 years. Current abstraction seems to be around 15–20 km3 yr1. During a couple of decades, the Saudi government had pumped several km3 yr1 of nonrenewable groundwater to grow low-cost crops (mainly cereals), which were heavily subsidized. The official aim of such an activity was to help transform nomadic groups into farmers. Now, the amount of groundwater abstraction has been dramatically reduced and the farmer nomads have become high-tech farmers growing cash crops. Another example is the situation of the Nubian sandstone aquifer located below the Western Desert of Egypt, where the fresh groundwater reserves are higher than 200 km3 and the maximum pumping projected is lower than 1 km3 yr1. Probably, similar situations exist in Libya and Algeria. Other examples of mining groundwater can be found in Llamas and Custodio (2003).

1.07.2.6 The Social Sustainability of Groundwater Management As previously stated, most aquifers present a large storage volume of groundwater in relation to their renewable resources (often two or three orders of magnitude higher). A practical consequence is that the potential problems do not usually become serious in the short term (within one or two generations). By then, the farmers may have experienced a positive social transition. Groundwater irrigation has proven to be an excellent catalyst for this social transition of farmers in arid and semiarid regions worldwide (Llamas and Martı´nez-Santos, 2005; Moench, 2003, 2007). Increased revenues result in, and allow for, a greater degree of social welfare. In addition, farmers are able to provide better education for their children, who may either move on to other economic sectors (generally more productive), or return to agriculture with a more productive outlook. Therefore, this transition means a reduction of global poverty (Lopez-Gunn and Llamas, 2008).

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Groundwater Management

This social transition, triggered by groundwater together with the implementation of more efficient irrigation technologies, can often result in a sustainable use in the midterm. However, adequate groundwater management and governance remain as important challenges in areas of India, Spain, China, or the southern United States (Shah et al., 2007; Llamas and Custodio, 2003; Foster et al., 2004). Aquifers constitute an example of common-pool resources, and as in the majority of cases, all actors have direct access (legal or illegal) to groundwater. Therefore, aquifers should typically follow the widely articulated tragedy-of-the-commons pattern (Hardin, 1968). Nevertheless, after half a century of intensive groundwater use, the authors of this chapter do not know any cases of medium-sized or large good aquifers (those with a surface larger than 500 km2, and medium-tohigh transmissivity and storage-capacity values) where the tragic outcomes outlined by Hardin have taken place causing social or economic disturbances – at least not in the degree of magnitude of those caused by soil waterlogging and salinization (India, Pakistan, or California), or the serious social conflicts in relation to people displaced or ousted by the construction of large dams (Briscoe, 2005; Shah et al., 2007). The situation may be different in small or poor aquifers, where storage is not large enough to sustain development for over two or three generations. Although still uncommon, cases of small aquifers that have run out of groundwater have been some times reported verbally to the authors of this chapter. The reality is that even some poor aquifers, such as the Indian hard-rock aquifers, have played a key role in increasing food production. In India, groundwater-irrigated surface has increased by more than 40 million hectares (ha) during the last few decades (Shah et al., 2007). As a consequence, India, despite an almost 100% increase of its population in the last 50 years, has not only achieved food security in practice, but has also become an important grain exporter. However, uncontrolled aquifer development in arid and semiarid regions worldwide raises sustainability concerns, particularly whenever the natural rate of recharge is low.

It might be appropriate to point out the situation of some large aquifers that have undergone overdrafting or groundwater mining for many decades. In many such areas, pumping data are hardly reliable. Take for instance, California’s aquifers, where overdraft estimates range between 1.2 and 2.4 km3 yr1. Equally, the overdraft in California aquifers has not been adequately analyzed since the 1980s. It is perhaps the lack of willingness to monitor, rather than overdraft per se that may constitute the greatest intergenerational threat for groundwater resources.

1.07.3 The Economics of Groundwater Use It is estimated that more than two-thirds of available freshwater is groundwater, and it is currently the most extracted natural resource in the world. More than half the world’s freshwater, for uses like drinking, cooking, and hygiene, comes from groundwater; groundwater irrigates 20% of irrigated agriculture. Groundwater supplies 75–90% of drinking-water supply in European countries, and 95% of the US rural population public-water supply. Aquifers provide natural storage reservoirs with little evaporative loss at little or no cost. Equally, aquifers provide natural transmission of water from the various sources to the point of use. During periods of drought, groundwater provides reliable supplies, compared to surface water, by its use as supplementary irrigation water to surface-water supplies (Howe, 2002). Groundwater is an important economic resource for billions of people, in developed and developing countries. Ninety percent of urban supply in India, and 70% in Mexico are just examples of the socioeconomic importance of this key resource for humans. Figure 4 plots the share of agricultural groundwater use and total groundwater use in total use in 2002 in the Organization for Economic Cooperation and Development (OECD) countries. It shows the importance of both the agricultural sector with groundwater use (countries situated on left) and the percentage of groundwater use over total use (countries on the right).

Po

rtu g G al re ec e N S et pa he in rla n M ds U ni ex te d ico St at e O s C D Ko E re Tu a rk ey EU -1 Ja 5 pa Ire n D lan en d m a Fr rk an Sw ce ed Be en lg G ium er m an Sl ov Au y ak s t U ni Re ria te pu d ki blic ng do Ic m el a C ze Hu nd ch ng R ary ep ub lic

100 90 80 70 60 50 40 30 20 10 0

% share of agriculture use in the total groundwater use

% share of total groundwater use in total water use

Figure 4 Share of agricultural groundwater use in total groundwater use, and total groundwater use in total water use. From OECD (2008) Environmental performance of OECD Agriculture since 1990, Paris, France. Online at: www.oecd.org/agriculture/env/indicators

Groundwater Management

Rosegrant et al. (2002) have estimated that the sustainable yield of groundwater resources in the world would be approximately 861 km3, down from 925 km3 as evaluated in the year 1995, with half this amount abstracted in Asia, and 28% in developed countries. It is difficult to put a dollar value to the use of this resource, but considering a conservative figure of US$ 0.25 m3 (including capital, environmental, and resource costs), this generates an annual total value of US$ 231 billion as a preliminary estimation. The groundwater literature shows that in addition to the direct-use value, groundwater resources also have a significant stabilization value (Tsur and Parker, 1997) in cases where groundwater is conjunctively used with more unreliable surface waters. Estimates of the stabilization value of groundwater resources show that it can be as high as that of direct use. This is because groundwater provides reliability of supply and reduces the probability and severity of water shortages. In addition to households’ water services, groundwater resources are extensively used for food production. With few exceptions, irrigation is the result of uncoordinated efforts of small entrepreneurs and farmers all over the world. These groundwater users have sought to improve their livelihoods investing in private capital and small pumping equipment to improve farm productivity.

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1.07.3.2 Productivity of Groundwater Use Despite the illusory accuracy of global irrigation data and the variability of the existing estimates, rough calculations yield the following conclusion: groundwater-based irrigation seems to be twice as efficient as surface-water irrigation in hydrological terms (m3 ha1), a ratio that increases to between 3 and 10 times from the social and economic points of view (US$ m3 and jobs m3). Regional-scale analyses carried out in Spain seem to confirm these figures (Herna´ndez-Mora et al., 2001) (see Table 1). Thus, it appears relevant and urgent to assess the comparative hydrological and socioeconomic efficiency of surface and groundwater irrigation at a global scale, carrying out similar studies in other regions of the world. Assessing the implications of this silent revolution should constitute a valuable contribution to the debate about global irrigation needs as perceived by many water experts. The required investment to assess the value and efficiency of groundwater irrigation versus surface-water irrigation can be afforded by most governments. Many high-value crops are watered with groundwater resources or by combining ground and surface water (Llamas and Martı´nez-Santos, 2005). For instance, in Table 1, Herna´ndez-Mora et al. (2001) show that, in Andalusia, irrigated agriculture using groundwater is economically over 5 times more productive and generates almost 3 times the employment than agriculture using surface water, per unit

1.07.3.1 Groundwater Costs of Abstraction and Groundwater Tariffs Groundwater unit volume costs increase with groundwater depth, as more energy is required for pumping and deeper wells might be needed. (In our experience, these costs usually range between US$ 0.02 and US$ 0.30 m3 depending on the country and the aquifer. However, according to Shah et al. (2007) the economic cost (value) of groundwater is about US$ 0.20–0.30 m3). It would be worthwhile to study this aspect worldwide in more detail since values appear very high in comparison to the general economic situation of Southeast Asia. One possible cause is the low technology used in the drilling of the wells and the performance of pumping devices. Groundwater irrigation cost per hectare also increases with time, albeit at a lower rate. This is because farmers begin to use a more efficient technology and switch (if soil and climate allow) to less-water-consuming crops: from maize or rice to grapes or olive trees, for instance. It is estimated that groundwater irrigation cost in Spain generally ranges between US$ 20 and US$ 1000 ha1 yr1. Despite the difficulties in setting tariffs for groundwater use, a number of countries have these in place. In many cases, tariffs are accompanied by quotas and licenses. In general, developed countries have in place a fixed fee plus a volumetric fee, but these levies are generally not adapted to recharge or movements in the water table. Essentially, this indicates that tariffs on groundwater use are environmental levies but not rationing instruments to manage aquifers. This is because, to ensure sustainable management, tariffs would need to be flexible enough to change according to scarcity costs and alluse externalities – and this – assuming that perfect monitoring and information are economically feasible.

Table 1 Comparing ground and surface-water irrigation productivity: Some irrigation economic indicators in Andalusia (Spain)

Irrigated area (ha) Average water consumption (m3 ha1) Total production (106 h) Production (h ha1) Employment generated (number jobs/100 ha) EU aid to income (% of production value) Gross water productivity (h m3) Total average water price to farmer (h m3)

Groundwater

Surface water

Total

244 190 (27%)

648 009 (73%)

893 009 (100%)

3900

5000

4700

2222

2268

4480

9100

3500

5100

23.2

12.6

15.4

5.6

20.8

13.4

2.35

0.70

1.08

7.2

3.3

3.9

From Herna´ndez-Mora N, Llamas MR, and Martı´nez-Cortina L (2001) Misconceptions in aquifer over-exploitation. Implications for water policy in southern Europe. In: Dosi C (ed.) Agricultural Use of Groundwater. Towards Integration between Agricultural Policy and Water Resources Management, pp. 107–125. Dordrecht: Kluwer.

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volume of water used. This difference can be attributed to several causes: the greater control and supply guarantee that groundwater provides, which in turn allows farmers to introduce more efficient irrigation techniques and more profitable crops. Greater dynamism has normally characterized farmers who sought out their own sources of water and have had to bear the full costs of drilling, pumping, and distribution. Higher financial costs to farmers motivate them to look for more profitable crops that will allow them to maximize their return on investments. Surface and groundwater distinctions, therefore, should be taken into account in order to achieve an efficient allocation of water resources.

The hidden or open subsidies that have traditionally been a part of large hydraulic projects for surface-water irrigation are probably the main cause of the pervasive neglect of groundwater problems among water managers and decision makers. Surface water for irrigation is usually given at low or heavily subsidized costs to farmers and this often results in the wasteful use of a valuable resource. It is usual that water supply companies, farmer unions, etc., lobby the state for the construction of surface-water infrastructures that are primarily paid for through general revenues, instead of advocating a responsible use of groundwater resources. At times, this may lead to social conflicts – such as in the case of the Tagus–Segura transfer or the overruled Ebro transfer, both in Spain – between water-importing and waterexporting basins. Progressive application of the user pays or full-cost-recovery principle of the European Union (EU) Water Framework Directive (WFD) would probably make most of the large hydraulic projects economically unsound. As a result, a more comprehensive look at water planning and management would be necessary and, in turn, adequate attention to groundwater planning, control, and management would probably follow.

came earlier. In this slowness and imperceptibility lies a significant part of the difficulties of managing groundwater resources nowadays. There are a number of political and policy choices to offset the pervasiveness of open access, such as regulatory interventions, market instruments, or information and technical choices. Most often, the effectiveness of policy measures might mean the right mix of policy instruments from an existing portfolio where experience has already been gathered from groundwater management. For policymakers, there are a range of policy instruments for management (Table 2), with different strengths and weaknesses for groundwater management according to a range of criteria such as effectiveness, economic efficiency, technical efficiency, administrative feasibility, equity, and social or political acceptability (Hellegers and Van Ierland, 2003). An example of regulatory instruments for environmental protection is the case of the Edwards aquifer, in Texas (USA). The Edwards aquifer is a karstic aquifer, which means that the effects of pumping are quickly transmitted to large areas creating in effect an open-access resource since it operates under the rule of capture, that is, where under Texas groundwater law, landowners can pump without limit (Howe, 2002). Pumping for agricultural and urban purposes has represented 45–50% of the total discharge of the Edwards aquifer for the period from 1934 to 1999. Springs supported several species of fish and amphibians giving the Federal Fish and Wildlife Service the right to intervene to protect these species if the state failed to act. A lawsuit by the Sierra Club, an environmental nongovernmental organization (NGO), under the Endangered Species Act ended in a federal ruling, which meant that the Texas legislature set up the Edwards Aquifer authority in 1996 with extensive powers, including the issuing of permits to regulate groundwater withdrawals. For example, it required pumping limits to protect endangered species. A flow of 150 cubic feet per second must be maintained at the most sensitive springs.

1.07.3.4 Groundwater: From Open Access to Common Pool?

1.07.3.5 Optimal Groundwater Pricing

Economics deals with scarcity and the allocation of scarce goods. Groundwater resources until very recently, it could be argued, were not economic goods, because anyone interested in pumping water could do it ab libitum. Capital, energy, or time constraints were the only barriers for all potential users of groundwater resources. In this case, however, the economic problem, if any, was related to access to inputs, finance, or labor. Even in cases where the inputs required to pump groundwater are unlimited, deciding on how much water should be pumped may not be an economic problem. One of the most pressing institutional questions centers on inappropriate legal and administrative structures, which in effect means that groundwater rather than being an openpool resource becomes an open-access resource which can lead to excessive contemporary and inter-temporal externalities (Howe, 2002) (see Section 1.07.6). As in many other natural resources, the transition from an open access and unlimited resource to an exhaustible and rival one is gradual. It happens by the marginal adhesion of small groundwater users attracted by the benefits that are obtained by those who

An optimal groundwater price is a theoretical economic concept, and the solution to a dynamic and stochastic problem. It is dynamic because the optimal price varies with time, and is meant to ensure that pumping rates at any given moment maximize the discounted flow of benefits for an infinite time horizon. As Neher (1990) shows, there is an optimal tax that makes all users of a common pool internalize the (marginal) consequences of their pumping rate so that a socially optimal management use is achieved. Theoretical optimal rates can be obtained to account for irreversible effects – caused by excessive drawdown (Rubio and Fisher, 1997), for situations in which backstop technologies can become operative if pumping becomes too expensive (Gemma and Tsur, 2007), or when surface water and groundwater are conjunctively used (Pongkijvorasin and Roumasset, 2007). Pongkijvorasin and Roumasset’s (2007) main contribution is to combine two problems that have been addressed separately in the literature. One looks at the problem of managing large surface systems with conveyance losses (Roumasset and Chakravorty, cited by Pongkijvorasin and Roumasset, 2007),

1.07.3.3 Perverse Subsidies in Water Policy

Groundwater Management Table 2

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Portfolio of policy instruments available to manage groundwater quantity and/or quality

Portfolio of groundwater-management tools

Example

Standards and regulations

High Plains Conservancy district No 1 outside Lubbock e.g., harim rule in Islamic law (Middle East) Indonesia EU Water Framework Directive For example, sprinkling bans to reduce low value agricultural groundwater abstraction in Holland; Yemen (ban on agricultural wells); and Egypt (ban on new wells) Netherlands, France, and parts of Germany have introduced a groundwater-abstraction tax Edwards aquifer Texas, (USA)a, Italy (Perugia province) Parts of Spain China

Limit pump capacity Well spacing Groundwater permits Safe-yield criteria Ban in most critical groundwater areas

Pumping tax, for example, reflect to the individual pumper the negative externalities Caps on abstraction Metering groundwater use Metering electricity use as groundwater use proxy Market-based instruments (economic incentives)

Administrative planning and education measures

Groundwater resource fee (money raised earmarked for aquifer management), cost of right (waterresource levy) Volumetric pricing of groundwater and/or electricity supply, sliding-scale pricing strategy Electricity subsidies Groundwater farms (i.e., purchase of land for associated groundwater rights) Issuing tradable permits (leasehold) with or without cap on total abstractions Water markets (freehold), for example, from rural to urban, from small to large farmers, from low-value to high-value crops, from irrigation to environmental flows Compensation program for third-party effects, for example, tax to area of origin Full cost recovery Groundwater banks Groundwater zoning, for example, aquifer-vulnerability maps to contamination Land-use planning

Education, for example, agricultural extension service and training Political education at senior level on the value of groundwater Public awareness campaigns, civil education, for example, value of groundwater Name-and-shame list Joint regulation and monitoring Participatory groundwater monitoring Institutional (incl. information and voluntary measures)

Incentives for private entrepreneurs for example, as franchisees for billing and collecting electricity dues Self-imposed correlative rights for example, % owned above the aquifer Voluntary agreements (self-regulation), devolved groundwater management Groundwater protection codes

Indonesia for industrial water users Holland has a levy under the 1983 Groundwater Act India and China For example, Mexico 1/3 of electricity costs Arizona Spain (Parque Nacional Tablas de Daimiel) Spain, for example, for environmental flows Spain, USA, Australia, Chile Mexico (including energy pricing coupled with water rights) Contemplated in Texas EU Water Framework Directive Trialled in Texasb United Kingdom, Holland, USA, and Canada Pakistan through informal committees For example, removal of invasive species and replacement with native vegetation in South Africa to help recharge shallow aquifers, India and South Africa studies on impact of forestry (native and plantations) on evaporation and recharge High Plains District Eastern La Mancha aquifer (Spain) e-Water India UNESCO For example, China water pollution map Eastern Mancha aquifer (Spain) India and China China

For example, Water Boards in Holland (interest, payment, authority) involved in groundwater-level management UK (Continued )

108 Table 2

Groundwater Management Continued

Portfolio of groundwater-management tools

Example

Technological

e.g. Spain and Mexico

Use of Geographic Information System (GIS) for monitoring Wastewater-reuse schemes Artificial aquifer recharge enhancement and/or storage Improved irrigation technology (for example, irrigation scheduling, micro-irrigation, land leveling) Changes in crop type, for example, higher-value crops if possible (i.e., no impact on livelihoods); ban on high water-consumptive crops

Central Valley (California), Rainwater harvesting (India); well recharge movement Israel, Spain, Australia, USA, and Mexico Areas in Spain Saudi Arabia

a

A system of marketable groundwater permits to be issued to all pumpers as a proportion to historical use subject to a total pumping cap. The cap was 450 mm3 yr1 which was accepted by all users, urban, agricultural and environmental. In 1998, a groundwater trust was set up to facilitate the trading of permits. In the period 1997–2001 there have been 403 trades with an average size of 235 000 m3. b For example, under the Irrigation Suspension program, water rights were purchased from 40 farmers representing 4000 ha to supply water to San Antonio. Irrigators were paid $ 98 to $ 1850 ha1 to stop irrigating, with water savings of 20 000 acre feet at a cost of $ 2–3 million paid for by cities, counties, and water companies. For additional ideas, please see Chevalking S, Knoop L, and Van Steenbergen F (2008) Ideas for Groundwater Management. Wageningen, The Netherlands: MetaMeta and IUCN.

and the other looks at the conjunctive use of surface and groundwater. The paper’s main accomplishment is to show that, with surface pricing as the only managing instrument, farmers switch from surface to groundwater and vice versa, when groundwater-scarcity rent diminishes, with the irrigation boundary contracting or growing depending on whether groundwater-scarcity rents increase or decrease. The institutional implications of these results are worrying, unless one can think of a much more restricting context in which pumping rights could be effectively enforced, and irrigation districts had nonmobile boundaries. Yet, Rijsberman (2004) quotes work by the International Water Management Institute (IWMI) which shows that groundwater use is largely beyond the possibilities of most water institutions in the developing world for a proper monitoring. Llamas et al. (2008) and Shah et al. (2007) present a similar general situation. The large literature on theoretical groundwater pricing yields somewhat impractical policy prescriptions (e.g., Molle and Berkoff (2006)) because of the number of externalities involved in many cases of intensively exploited aquifers (see Llamas and Custodio, 2003). Brown (2000) offers very convincing arguments to explain why optimal pricing has rarely been used to allocate renewable resources such as fisheries or groundwater.

1.07.3.6 Departures from Optimality: Second-Best Solutions In economics, second-best solutions are those that cannot achieve the results of the first-best optimal solution but are more applicable in practical terms. First-best solutions may be too information demanding or based on a perfect fine-tuning to the specific circumstances. In this section, we review some of the most commonly used second-best (or even third-best) solutions to properly manage groundwater resources. These policy approaches are always implemented to solve one or the other type of the sources of economic inefficiencies. The following are the policies from the less to the more sophisticated: user rights, pumping rights, pumping quotas, water tariffs, and water markets.

Issuing user rights is the simplest way to grant access to an aquifer. The authority may or may not control the pumping capacity and the type of equipment. In intensively exploited aquifers, granting user rights may not be sufficient to deter pumping and, in many situations, the outcome may not ensure that the exploitation is adequate. However, issuing user rights is a prerequisite to consider in any of the policy menus mentioned. In principle, user rights are increasingly required to obtain legal access to tap an aquifer in virtually all contexts in which water is scarce. The next instrument in the list is granting pumping rights, whereby users are entitled to pump fixed amounts. In principle, the ownership of pumping rights can be associated with private property if those rights cannot be encumbered by new users or forfeited by a public agency. Setting up pumping quotas enables a more flexible instrument, because these can be modulated to the aquifer’s recharge. In many cases, pumping rights are combined with water tariffs, but an interesting option would be to modulate tariffs to mimic the optimal price. Perhaps, the most advanced initiatives in the area of groundwater pricing can be found in some European countries, which have seized the opportunity of the EU WFD to implement the principle of full-cost-recovery pricing. Water markets can be found in multiple formats. In India, for instance, water markets occur as informal exchanges among small farmers (Saleth, 1996). In Australia, water markets are formally established and can facilitate groundwater trading or rights’ exchanges. In Spain, the basin authorities have offered farmers permanent buy-outs of water rights, or annual pumping quotas, following public offerings (see Box 1 for examples of water trading). A prerequisite of all these policies is to have control of (at least) the number of users, and ideally of the pumping yields of all users tapping an aquifer. However, even if control and surveillance can be guaranteed, it does not imply that the aquifer will be sustainably managed. Unfettered water rights can be as damaging to aquifer management as an aquifer that is not controlled. Groundwater management is usually based on rights or licensing, but enforcing compliance with these

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Box 1 Water markets and voluntary exchange mechanisms to solve groundwater overdraft. From Garrido A and Llamas MR (2009) Water management in Spain: An example of changing paradigms. In: Dinar A and Albiac J (eds.) Policy and Strategic Behavior in Water Resource Management, pp. 125–146. London: Earthscan. Water banks or exchange centers, as these are called in Spain, received legal recognition in the 1999 Spanish Water Law reform. Not strictly a bank or agency, these centers are hosted, run, and located in the basin agencies themselves. It is widely believed that these centers are a much more efficient medium for promoting water exchanges, for a number of reasons, such as transparency, control, avoidance of third-party effects, and market activity and scope. Yet, the experience so far has been limited to the Ju´car, Segura, and Guadiana basins, since these water centers have been primarily used to tackle severe problems of groundwater intensive use. Since the enactment of the 1985 Water Law, which included special provisions to tackle the problem of overexploited aquifers, there have been at least four major initiatives to manage groundwater resources. In short, these were (1) the declaration of overexploited aquifers and the mandate to enforce regulations and implement management plans; (2) an EU agri-environmental program, only applicable to Aquifer 23 in the Guadiana Basin, with subsidies to farmers who curtail their water consumption; (3) the use of inter-basin transfers, both in the case of the southeast coastal areas and in the Upper Guadiana; and lastly, (4) the Special Upper Guadiana Plan (PEAG, Spanish acronym), and the creation of exchange centers in the Segura, Ju´car, and Guadiana basins. Clearly, the first option failed; the second one succeeded, but the financial cost was very high, and the third option failed because the second one was not sustainable. In the end, the PEAG was approved in 2007 with a total budget for 20 years of h5.5 billion (equivalent to the proposed Ebro transfer) and part of its subprograms are now operational, although under PEAG the basin would reduce to a meager 200 million m3. Underlying these initiatives, but undermining them too, was the recognition that tens of thousands of users in virtually all basins had no legal rights or concessions to the groundwater resources they had been tapping for years. Any effort to reduce total extractions in the over-drafted hydrogeological units had to be accompanied by the closure of the alegal or illegal uses. In 2005, it was clear to all managers, analysts, and users that something new had to be given a chance. The option to use buyouts of water rights, permanent or temporary, gave a rationale to the establishment of exchanges centers (centros de intercambio in Spanish). We review the different approaches taken in the Jucar and Guadiana. In the Jucar basin, the offer of public purchase (Oferta pu´blica de adquisicio´n de derechos, OPA) was targeted at farmers tapping groundwater resources near the Jucar’s headwaters. Its objective was to increase the water tables in Castille-La Mancha to ensure that the Ju´car flows to the Valencia region increase from historical lows. Farmers were given the option to lease out their rights for 1 year in return for a compensation ranging from 0.13 to 0.19 cents m3, the variation depending on the distance of the farmer’s location to associated wetlands or to the river alluvial plain. The OPA was launched in two rounds, the first with disappointing results in terms of farmers’ response, while the second had more success. The purchased waters served the unique purpose of increasing the flows, enabling more use downstream in Valencia. However, the OPA did not have any specific beneficiaries downstream, other than to increase flows. The OPAs of the Guadiana followed a completely different approach and were meant to address serious problems of overdraft in the Upper Guadiana. As stated before, the OPA formed part of the more ambitious program of aquifer recovery, the PEAG. The Guadiana’s OPA made offers to purchase permanent water rights to groundwater, paying farmers h6000–10 000 ha1 of irrigated land. Note that, since these farmers had seen their allotments reduced in preceding years, what the Guadiana basin was truly purchasing from the farmers was about 1500–2500 m3 ha1, effectively h2–4 m3. The Guadiana basin agency has the objective of purchasing the water rights of 50 000 ha1 of irrigated land, and is budgeting h500 million for the whole plan. A marked difference from the Jucar’s OPA is that the Guadiana exchange center will transfer part of these rights to other farmers (growing vegetables) and to the autonomous community of Castille-La Mancha. The Guadiana basin will grant less rights than it has purchased, allocating the difference to wetlands and to increasing the piezometric levels of the aquifers.

rights remains a major challenge for adequate groundwater management. For example, there is wide experience with water markets in the USA; according to Howe (2002), what water markets do best is to generate information on values for more rational, better-informed water allocation – for example, the sale or lease of water rights to off-site buyers such as cities. In Arizona, the government had acquired 200 000 ha of land by 1990 for the associated groundwater rights. These water farms or water ranches average about 12150 ha and are valued at US$15 million, and expected to supply 15 000 acre feet of groundwater per year for 100 years. In Colorado, Front Range water rights in 1990 sold for US$ 1000–4000 per acre foot (Colby, 1990 in Wagner, 2005). In systems where groundwater rights have been incorporated into the general water-rights system, groundwater rights can be bought and sold, transferred to other locations, or transformed into surface rights (e.g., tributary groundwater– groundwater that is intimately connected with surface water). In Texas, groundwater was purchased to secure water for urban centers such as Houston, San Antonio, and El Paso. For example, an old mining right from the Alcoa-Sandown mine was sold for US$ 688 per acre foot annually. The city of Amarillo also bought groundwater for US$ 679 ha1 for groundwater rights from 28 350 ha of lands, when the land itself sells for US$ 494 (Gillinland, 2004 in Wagner, 2005). The El Paso Water Utility purchased more than 19 000 ha of ranchland to pump 15 000 acre feet by 2010 (Texas Center for Policy Studies

in Wagner, 2005). Lucrative groundwater leases with at least four private water ranches on over 200 000 ha have been formed to sell or lease a significant amount of water to off-site users, principally cities. Another example is the Irrigation Suspension Program, where water rights were purchased from 40 farmers representing 4000 ha to supply water to San Antonio. Irrigators were paid US$ 98–1850 ha1 not to irrigate, with water savings of 20 000 acre feet at a cost of US$ 2–3 million paid for by cities, counties, and water companies (Wagner, 2005). Nevertheless, there is a range of issues that has to be considered in water marketing. Over-entitlement occurs when the sum of all legally defined rights are greater than 100% of the system’s potential. Overuse occurs when the quantity of water abstracted is greater than the system’s potential to supply. Sleeper or dozer licenses are legal entitlements that are not used at present but are legal (e.g., UK sleeper licenses). One major reason why some people might distrust markets is the fear that markets might fail to deal with issues of distributional equity, fairness, public concern, and community interest.

1.07.3.7 Internalizing the Value of Environmental Services Provided by Groundwater Water can be framed on current discussions on ecosystem functions from environmental services. Ecosystem functions refer to system properties and processes. Services represent the

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benefits that society derives, directly or indirectly, from ecosystem functions. A summary of the authors’ evaluation of annual flows of water-related ecosystems at the world scale is presented. Humans avail many types of services from water-related ecosystems in addition to water supply. Note, for example, that 1 ha of wetlands can generate almost US$ 4200 yr1 in wastetreatment services (Costanza and de Groot, 1997). While this evaluation was certainly preliminary at the time it was produced, it conveys a clear idea about the costs and damages that water scarcity can provoke. The mere recognition of many of the identified services valuable for society has huge implications for drought-policy design and implementation. Chief among this is the fact that many of these services have public-nature features, which means that they are nonrival and nonexclusive goods. As scientists have learned to identify and value them, water policy must take into account and ensure that decisions are a compromise between both productive and nonproductive services (National Research Council, 2004). The Millennium Ecosystem Assessment undertaken in 2005 classified the goods and services provided by natural resources as provisioning, regulating, supporting, and cultural services. In this context, Bergkamp and Cross (2007) discuss the high-value ecosystems supported by groundwater. The total economic value (TEV) of groundwater resources is the sum of groundwater resources: use and non-use value, based on direct values, indirect values, option, and existence values. A number of methods have been developed by environmental and ecological economists to value these goods and services using a range of environmental techniques such as contingent valuation method (willingness to pay and willingness to accept), choice modeling, production-function approaches, surrogate markets, costs-based approaches, and stated preference. These methodologies are addressed in one of the chapters of this volume. This is a way of internalizing the value of groundwater services and also the economic value of groundwater externalities. This can help decision making to evaluate the costs of action and costs of inaction. It gives a clearer signal of whether it is better or not to opt for preventative policies when the full remediation costs of polluted groundwater are accounted for, that is, measure the potential benefits (or damages) of a range of effects. In 1996, the US Department of Defense invested in 75 pump-and-treat systems to remediate contaminated sites estimated at US$ 500 000 per site. Equally, a calculation on the externality costs of overabstraction in the Queretaro aquifer (Mexico) between 1970 and 1996 included: the increased pumping costs due to drop in piezometric levels estimated at US$ 6 million; the loss of water quality for public water supply at US$ 26 million; and damage to urban infrastructure due to land subsidence at US$ 26 million (6 million as private cost and 20 million as costs to taxpayers). Moreover, it is relevant to consider impacts on other policy sectors like the costs for public health due to the mobilization of naturally occurring arsenic through deep wells in Bangladesh, which is estimated to affect 30–35 million people (WHO, 2001). Opportunity costs are the foregone benefits that could have been generated if resource was allocated to the next-best use if water is not allocated to its highest use value and, in fact,

opportunity costs may be greater than the value generated by next-best use. In this case, the economy is subjected to inefficient and suboptimal groundwater allocation, although this may be justified in equity or sociopolitical terms. Therefore, it is equally important to properly include the positive services provided by groundwater. In terms of equity, Acharya and Barbier (2000) analyze losses to farmers from reduced groundwater recharge. On average, farmers could lose US$ 413 ha1 if the groundwater benefits are not accounted for, that is, evaluating the systemwide benefits associated with groundwater use. The increased recognition of the benefits associated with groundwater use is reflected, for example, in the growing aquifer-recharge movement. In Texas, aquifer recharge through open-space protection and cooperative groundwater allocation is a new paradigm in water management (Wagner, 2005), based on valuing the products and services that a functioning system provides. In India, meanwhile, the socalled decentralized recharge movement was a spontaneous response to groundwater depletion to help water tables rebound to predevelopment levels at the end of the monsoon season in pockets of intensive use. This is an example of contrasts between popular hydrogeology and formal hydrogeology; for example, scientists argue that hard-rock areas have too little storage and advocate recharge; meanwhile, the prolific growth of recharge structures is based on the value people attach to a check-dam even if their wells provide only 1000 m3 which – although small – is crucial for life-saving irrigation in times of delayed rain. Thus, rainwater harvesting can be used to recharge groundwater via recharge ponds, based on the main sources of recharge: rain, and infiltration from riverbeds and from the floodplain. Equally, in Australia and the USA, sand dams are used to make artificial shallow aquifers in streambeds to reduce evaporation of stored waters. The growth in the number of sand dams could substantially increase to compensate for potential climate variability and change making use of groundwaters’ buffering service (see Section 1.07.6.2.1). Failing to account for the opportunity costs of groundwater and surface/groundwater linkage often result in suboptimal outcomes. Groundwater recharge is one of the most important environmental functions brought about by wetlands. For example, households in rural Nigeria rely on groundwater for drinking and cooking, and in the arid north, particularly during drought, it often has the added advantage of its higher quality (Table 3).

1.07.4 Regulatory Frameworks for Groundwater Multilevel Governance One of the most obvious examples of the Cinderella status of groundwater in global water resources is reflected in the evolution of regulatory frameworks. Due to its silent (and relatively recent) rapid growth, groundwater traditionally had little or no regulation (i.e., as exemplified in the rule of capture in Texas), part of mining law, or of private-property rights (tied to land). This lack of prominence and the lack of concern over its management and state of preservation have historically been reflected in the law. Therefore, groundwater

Groundwater Management Table 3

111

Valuing groundwater goods and services

Total economic value of groundwater

Environmental goods and services

Definition

Example

Direct value

Provisioning service

Drinking water supply

Two billion people rely on groundwater directly for drinking water More than 50% of cities with population of more than 10 million rely on or make significant use of groundwater Forty percent of world’s food relies heavily on groundwater Land irrigated from aquifers has increased 113 times between 1990 and 1990. Aquifer supplies more than half the world’s irrigated land For example, manufacturing processes and geothermal and cooling systems Groundwater stores and releases water, sustains river flows, springs, and wetlands Through microbial degradation of organic compounds and potential human pathogens, microbiological and some chemical contaminants removed, retarded, or fragmented For example, by absorbing run off Primary buffer against climate variability and spatial variability of droughts Potential innovation as future use for anthropogenic carbon sequestration in the ground For example, groundwater recharge and discharge

For example, storage and retention Agriculture

Industrial Indirect value

Regulating services

Water regulation; Water purification and waste treatment Erosion and flood control Climate regulation

Option value and existence value

Supporting services

Cultural services

Necessary for the production of all other ecosystems services Non-material benefits people obtain from ecosystem services

Spiritual enrichment, cognitive development, religious value, and symbolism

From Bergkamp G and Cross K (2007) Groundwater and ecosystem services: Options for their sustainable use. In: Ragone S, de la Hera A, Hernandez-Mora N, Bergkamp G, and McKay J (eds.) Global Importance of Groundwater in the 21st Century: The International Symposium in Groundwater Sustainability, pp. 233–246. Alicante, Spain, 24–27 January 2006. Westerville, OH: National Groundwater Association Press.

was legally structured as one more facet of the right of ownership for a specific area of land. Starting out from that premise, the various laws gave shape to the depth of that right and regulated how it would fit in with the rights held by owners of adjacent pieces of land. All of the above emerged from an eminently private perspective imbued with the wealth of duties and rights conferred by ownership rights. Groundwater doctrine in Texas is based on the rule of capture, an English common-law approach based on absolute ownership, where landowners can pump without limit, as long as water is put to beneficial use, which allows unrestricted pumping by competing groundwater users as long as it is not wasted, and whereby property rights are not defined (Wagner, 2005). This is an example of one of the few natural resources in the USA not regulated by a central agency. The activities of 88 water districts in Texas are unusual because they have been based exclusively on a voluntary approach controlling wastage of water, recharge, enhancement, and water-conservation education rather than controlling abstractions. For nearly 100 years, the rule of capture has survived attempts to regulate groundwater use. Although government oversight and technical assistance are vital, a carefully crafted free-market system based on private rights to a communal resource becomes increasingly important. A bottom-up process created the State Water Plan of 2002, which incorporated regional water plans’

gradual increase in the security of mining rights from open access to other systems.

1.07.4.1 Diversity in Groundwater Regulatory Regimes Regimes with a civil-law tradition, inherited from the principles of Roman law, are clear exponents of this and not very far from them are those grouped under the parameters of common law. Under common-law regimes, the landowners were the right holders of groundwater, flowing under their properties, which could be harnessed (Embid Irujo, 2002). From the legal viewpoint, legislation on aquifers presents two main issues of concern: first, ownership and second, transferability or flexibility with ownership rights. The first one relates to whether groundwater resources should be public or private property. Ownership of groundwater resources shows a high level of diversity from completely private (e.g., in Texas USA), groundwater from the Ogallala aquifer is mainly private (Peck, 2007), to state-owned resources, such as in the case of Mexico, to plural legal systems, as in parts of Africa, and community based, such as in Bolivia. Legal provisions may confer ownership of groundwater directly on public authorities, as part of the public domain (Morocco, Italy, Spain, Zimbabwe, Israel, some US States, Jamaica, Mexico, Argentina, Australia, the Lebanon, Jordan, and Syria), give the authorities preferential rights in groundwater (South Africa and Uganda),

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or make groundwater the common heritage of the nation and place conditions on its use (France). Particularly, in Muslim countries, the applicable legal regime is intertwined with principles related to religious beliefs which result in the classification of water as public in itself with cases of private appropriation being seen as exceptional (Caponera and Nanni, 2007). Whichever route is taken, the result is very similar: governmental authorities give themselves powers over groundwater with the aim and effect to implement public policies that they lay down. The progressive importance of groundwater in the definition of water resources in various countries has engendered the implementation of legal reforms that protect public intervention (Hodgson, 2006). When groundwater is public, the concept that is generally used by the rule makers is the permit, license, authorization, concession, or a similar instrument. This is the case of Israel, a number of states of the USA, Mexico, and many other countries. In other places, such as California, Chile, India, or Texas, groundwater is under private ownership. In all of these cases, a private party, individual, or community is granted the right to use a certain amount of groundwater. This right is subject to certain conditions relating to time or use. The right may or may not be granted according to whether it is consistent with the status of the resource and with the parameters of the planning regulations on water resources that must govern its contents. Those same premises must be used to determine the period for which the right is granted and the amount that may be extracted. It is also necessary for the law to outline the cases in which this right can be altered, restricted, or even eliminated as a result of damage to the aquifer, possible droughts, watersupply needs to the population, or similar events, or principles such as reasonable and beneficial use. A legal regime for groundwater should ideally consider some of these aspects:

• • • • • •

•

approval of compulsory legal norms for all groundwater users; determination of the legal rules and principles to be applied in the management of groundwater, including its relationship with surface water; legal parameters to define groundwater as a resource; institutional regime applicable to groundwater; specification and regime for uses of groundwater available for all citizens without being subject to specific control; determination of the rights on water (transitional rules in the event of amendments to the law; concession rules; registration processes; contents of the right – volume of water, term, conditions, and termination – transfer of rights; and dispute-resolution mechanisms); and rules on the protection of groundwater and measures to adopt if needed (control of the pollution of groundwater – rules on discharge and authorization; use restrictions; and prohibition on use).

1.07.4.1.1 The controversy over private, public, or community groundwater rights Evidence indicates that ownership per se (public, private, or common) does not guarantee or pre-empt sound management. The importance does not lie in what name is given to the legal title but rather in the contents given to that title. The

emphasis is placed on the aim to be achieved. Preferences on ownership are societal choices, which are subject to change and flux; the underpinning question is not ownership but whether management is according to some predefined a priori objectives, which in any case are themselves subject to constant negotiation and renegotiation as part of a normal political process. Some authors consider that the legal declaration of groundwater as a public domain is a conditio sine qua non to perform a sustainable or acceptable groundwater management. This assumption is far from evident. For many decades, groundwater has been a public domain in a good number of countries. Nevertheless, sustainable groundwater management continues to be a significant challenge in many of those countries. Highly centralized management of groundwater resources is not automatically the solution to promote solidarity in groundwater use as a common good because a key element is the internalization, by often thousands or hundreds of individual users, on the need for collective action. Groundwater management sometimes can successfully be devolved to stakeholders of the aquifer, in self-governance arrangements under the supervision of the corresponding water authority. Stakeholders’ participation has greater chances of success if it emerges bottom -up and is supported topdown. The practical application of a hybrid (public and private) system is exemplified by a few countries in the world where a range of systems coexist: one is the US, where states like Colorado, Arizona, Texas, and California exhibit a range of ownership rights to groundwater; and the other is Spain with a particularly interesting example of a mixed system. Wells drilled after 1 January 1986 require governmental permission, while those operational before 1986 remain private. Private groundwater may remain so for 50 years (provided the well owners reach an agreement with the government in exchange for administrative protection) or perpetually (if the owner wishes to preserve his/her rights under the 1879 Water Act). In any case, the Spanish situation is far more complex due to the lack of a reliable registry of groundwater rights. While the government is currently carrying out a series of remedial initiatives, these ignore a significant share of existing wells, and the registry or inventory is therefore incomplete. A key ingredient is the need for a strong political willingness to apply the laws. It seems clear that a reliable inventory of groundwater rights is desirable in order to ensure adequate management. The second issue refers to the way groundwater rights should be inventoried and to whether the possibility to trade with them should be allowed. This second aspect, usually equated with water markets and banks (discussed in the previous section) is perhaps subordinated to the first in terms of importance, even if significant informal markets already exist in some places (Mukherji, 2006). It cannot be ignored, however, that in other territories, the inseparable link between water, land, and private property has been maintained. The states of California and Texas or countries such as Chile and India are examples of this. These cases have maintained the private ownership of water and in many cases applied the doctrine of prior appropriation, although this is subject, depending on the territory and circumstances, to specific measures of administrative or court intervention, linked, for example, to the principle of reasonable use. It is also possible

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Box 2 Groundwater-use rights in China. By Simon Howarth, based in Gansu, under the UK Department for International Development (DFID)-funded Water Resources Demand Management Assistance Project. The Shiyang River Basin in Gansu province in Northwestern China is an area of severe water shortage, where groundwater abstraction greatly exceeds recharge. A package of measures, including greater delivery of surface water and restrictions on the use of surface water, is being introduced. A key element in this process is the allocation of water rights to individual households. These are based on land allocations and household size (to take account of domestic use and livestock as well as agricultural demands). Both land and water are owned by the state in China, but user rights have been granted to individuals. Land rights have been granted since the 1980s and further reforms are in progress. Water rights have been formalized since 2007, when individual household water-rights certificates were issued (via village committees or WUAs). The rights are calculated to be sufficient for the locally recommended cropping pattern, and are being reduced each year (e.g., from 7200 to 6615 to 6435 m3 ha1 in one typical irrigation district) as recommendations are revised and farmer skills in water savings are developed. Awareness and capacity-building programs are being run simultaneously so that farmers can protect their livelihoods while coping with less water. Wells were developed by villages and are owned by them, but the amount of water that can be pumped from them is regulated by the state through a system of permits. Well permits are now being reissued to suit the new water rights, and electronic controls (IC cards) are being installed at all wells. These will limit the amount that can be pumped from the well to the annual total permitted for that well. These cards are held by the well operator who is responsible for ensuring that each household receives water in accordance with their individual rights. These systems are new, rely on both sophisticated technology and complex administrative systems, and have been introduced rapidly (in over 10 000 wells in 1 year). Not surprisingly, some teething difficulties have been encountered, but there is a very strong political will to solve these problems. Allocation of water rights is intended to enable trading of rights, although this does not yet happen on a formal basis in this area of China. It will be subject to certain restrictions – for example, the right will be salable at a maximum of three times the water-resources fee, which is small when compared to both the value of water and the cost of pumping. There is a large and growing requirement for water for industrial development, which is a more valuable use of water, but there is a competing requirement for food security – these competing demands cannot be managed purely by market measures but will require government control as well.

to outline mechanisms for exchanging water rights in the context of what has come to be called the water market (Chile, South Africa, Mexico, Spain, the UK, Australia, and the US). The use of water-market institutions for groundwater has its detractors who point to the risks involved. In our view, however, those risks do not necessarily warrant ruling out the option completely. Bringing flexibility to the allocation of resources and allowing them to be exchanged are not in themselves misguided concepts. The usefulness of water markets is usually associated with cases of multiple supply and demand sources, with transparent exchange mechanisms and the appropriate transportation networks that make them feasible (Melgarejo and Molina, 2005). The key lies in controlling their use and making them subject parameters of sustainability and protection (Box 2).

1.07.4.2 Implementation and Enforcement of Groundwater Legislation In the context of groundwater management, rules on the ground are crucial, for example, those related to time, well location and spacing, technology, or groundwater-abstraction quotas. In addition, another factor is the interaction between formal groundwater law and its operation on the ground. More attention is being paid increasingly to the implementability of regulation, since the problem with most groundwater legislation lies in its implementation and enforceability. For example, South Africa established an implementation team with the task of anticipating what the water domestic bill would require, with close interaction between the drafting and implementation teams to identify possible implementation problems before enactment. In the case of groundwater, it would be useful to develop implementation tools such as guidelines, procedures, information systems, user manuals, and organizational arrangements. Another option is to opt for framework laws, which specify general guidelines but leave implementation to detailed regulations as used in Uruguay.

Implementation requires time, and needs political support at the highest level since strong economic and political interests are usually affected by allocating or reallocating groundwater resources. As Gardun˜o (2003) states that implementable legislation is one that the government is able to administer and enforce and water users have the ability to comply with. Experience shows the education of stakeholders and widespread presence of groundwater-user associations is crucial for an adequate participatory bottom-up management approach. One of the main problems in groundwater governance is lack of enforcement in some cases of relatively sophisticated laws, such as in Spain. As stated earlier, institutions encompass not only rules in norm but also rules in use or institutional arrangements. In effect, the implementation and enforcement of groundwater laws have to be legitimized and supported by society (social norms). The involvement of groundwater users in groundwater-management regimes is a necessary (although not sufficient) condition for successful enforcement regimes. In traditional societies, social networks were denser and therefore transaction costs lower, whereas modern societies require complex institutional structures that constrain and regulate interactions among groundwater users. Groundwater users possess detailed local knowledge on water use, and these communities can apply for sanctions unavailable through formal institutions. For example, name and shame can resolve conflicts at the local level in a manner customized to local circumstances, which reduces transaction costs, which in turn are critical for economic performance. In fact, in a study undertaken in Spain, groundwater users had a clear perception of the kind of behavior that should be penalized and how sometimes sanctions devised by farmers do not mirror sanctions designed by higher-level authorities (Lopez-Gunn, 2003). This can be rooted in different perceptions of equity and fairness. For example, farmers in an aquifer in Spain would prefer to be sanctioned in the following irrigation season with water as a penalty, in lieu for the same amount of water that farmers abstract over their quota in the previous

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irrigation season, instead of the current (formal) sanctions of a monetary penalty. Thus, groundwater users can reduce the transaction costs of enforcement and devise adequate sanctions. However, it should not be forgotten that authorities in most cases ultimately hold this legal responsibility to protect public goods. Higher-level authorities will often have to be imaginative with monitoring and sanctioning regimes. Many Asian administrations lack the capacity to perform complex tasks, for example, urban groundwater management with joint monitoring of industrial groundwater abstraction and wastewater discharges. A feasible alternative tried in Indonesia was to select a random sample and thoroughly monitor these users. In cases of noncompliance, the weight of the law should be applied and widely publicized in the media; as capacity grows, the sample could be enlarged. Limited administrative capacity is a key constraint to groundwater management and revenue-raising fees can be re-invested toward capacity programs.

1.07.4.3 Multilevel Regulatory Frameworks The section above described and discussed briefly some of the main challenges for national groundwater law. It is increasingly recognized, however, that national groundwater law is only part of the regulatory framework. Other levels (both conceptual and in terms of scale) have to be taken into account: first, international conventions currently being negotiated, for example, for transboundary aquifers or the rise in the human right to water; second, a pragmatic approach on the advantages and limitations of legislation and litigation; and third, a consideration of the legal principles that have to underpin legal norms and an evolution in our understanding of how laws will be drafted in the twenty-first century. First, in relation to international conventions, there are two conventions that are applicable to groundwater: the first relates to transboundary aquifers and the second refers to the International Convention on Human Rights (1948) and its new impetus to recognize a human right to water (or HRW). Until only a few years ago, international law did not pay too much attention to groundwater. This state of affairs has changed, aided by the Convention on the Law of the NonNavigational Uses of International Watercourses (1997) (Eckstein, 2004). This Convention, yet to be ratified, only partially covered transboundary groundwater, that is, those connected to rivers, and thus left many aquifers uncovered. As a result of this situation, in 2008, the International Law Commission delivered to the United Nations General Assembly, draft articles for the law on transboundary aquifers. After reaffirming the protective and environmental approach to the use of groundwater, they ratified the application of the principle of fair use (1997) and of sensible damage. They also outlined measures for the following: first, cooperation between states; second, the regular exchange of data and information; third, the promotion of bilateral and regional agreements; and fourth, measures for the protection and preservation of ecosystems, and the prevention, reduction, and control of pollution. Along these lines, they provided that where appropriate, a shared management mechanism will be established.

Claims for the right to water to become a fundamental right, and thus protected, are increasing. This is probably highly applicable to groundwater since in many countries public water supply (to which the HRW is addressed) is supplied largely by groundwater. This is the case, for example, in Africa, the continent lagging most behind in the Millennium Development Goals. In relation to lack of access to water and sanitation by 2015, Africa is the continent most off target where groundwater is the daily source of drinking water for more than 75% of the population The first reference in this respect is to be found in articles 11 and 12 of the International Covenant on Economic, Social, and Cultural Rights of 19 December 1966. While not expressly mentioning the right to water, its wording has led the United Nations Committee on Economic, Social, and Cultural Rights (2002) to define the HRW as one which entitles everyone to sufficient, safe, acceptable, physically accessible, and affordable water for personal and domestic uses, and even links this right to the International Bill of Human Rights (1948). This reference to the right to water has been kept in recent documents such as the Plan of Implementation of the World Summit on Sustainable Development (2002), the Charter of Water of the Senegal River (Mali et al., 2002), or the Third World Water Forum Ministerial Declaration (2003). Second, there are advantages and disadvantages to a pure regulatory approach. Recognized limitations include symptoms such as the existence of rigid overly bureaucratic administrative procedures, the large number of authorities involved in taking decisions on groundwater, scarcity of technical and human resources to enforce compliance with legislative requirements, the often-absent citizens’ participation in decision-making processes on deliberation and decision making, or the confrontation of interests among the various government departments. These are clear examples of what we could call organization sickness. Legal proceedings that are prolonged, costly, hard to enforce, or construed poorly with practical needs of water management make it problematic for courts to be able to solve groundwater conflicts. Crucial and fundamental advantages to regulatory processes remain, such as its role as leverage and recourse for aggrieved third parties in court. This is why it is crucial in the case of groundwater to facilitate legal literacy or legal empowerment improving the capacity of communities to know and use the law – training in techniques such as interest-based negotiation, mediation, and facilitation. Third, the step from regulation by rulemaking process to negotiated rulemaking can never replace the public decisionmaking process, with the participation of all interested parties, or generate inequality. A series of legal principles have to be embodied in formal groundwater regulation, leaving more freedom or flexibility in terms of the implementation and enforcement (Table 4).

1.07.5 Institutional Aspects of Groundwater Management The problem of groundwater over-use has often been portrayed by the tragedy of the commons, that is, Hardin’s seminal essay in 1968 (Hardin, 1968) which describes how the

Groundwater Management Table 4

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Legal principles applicable to groundwater legislation and its implementation

Legal principles

Rationale and justification

How

Effectiveness and efficacy

Efficacy of water management must be sought, by implementing or furthering measures

Cooperation

Participation and subsidiarity

Cooperation between authorities as fundamental (Declaration on Groundwater in the Mediterranean, 2006)a Aarhus Convention (1998)b

Sustainability and precautionary

Rio and Johannesburg Summitsc

Common responsibility

Commission on Sustainable Development, United Nations Economic and Social Council (2008)d

Adapting organization and competent authorities to conform to the natural characteristics of the resource – normally identified as a drainage basin Fostering the participation of users and interested third parties which has already been identified as a mechanism to secure acceptance and implementation of the agreed measures. Encouraging planning related to the allocation of resources or water-quality protection or restriction measures, rules on improvements and irrigation transformations, guidelines on recharge and aquifer protection. Cooperation, either through procedures or by agreeing to specific conventions, must enable more effective, allow the views of each of the players with responsibilities in the area to be known, avoid subsequent defects in the implementation of agreements and, in short, allow views to be joined to find the best solution. Environmental governance that is transparent, legitimate, and efficient. Public authorities, as the necessary guardians of correct application of the legal framework, may confer an especially important role on user associations directly involved in the management of, for example, groundwater resources. There has already been a certain amount of international experience in this area in countries such as Argentina, Colombia, Spain, the US, Indonesia, Mexico, Nepal, the Philippines, Sri Lanka, or Tunisia. Besides, it acquires greater importance in relation to groundwater as it is a way of surmounting the management difficulties caused by having multiple users. The implementation of sustainable development must pervade decisions on territorial and urban planning and the performance of specific projects, the approval of new protection rules, to end, cease or modify granted rights to groundwater and, especially, the economic development and growth initiatives in every country. Groundwater is a common good and therefore the responsibility for its protection and correct management belongs to everyone.

User and polluter pays principle

Solidarity

Levels of solidarity in groundwater management

To determine the obligation to repair and replace the resource base to their original state. In addition, it will be absolutely necessary to establish the strict liability regime in these cases, notwithstanding any potential exceptions linked to the state of technology or the grant of approvals.  Intergenerational solidarity. Future generations must be considered when adopting initiatives.  International solidarity. Not all countries have the same difficulties. Ranging from the actual exchange of water to the transmission of technology and knowledge. An example is the Johannesburg Declaration on Sustainable Development of 2002.  Regional solidarity. The areas within a state must seek points of consensus and foster instruments of cooperation in the rational and sustainable use of groundwater. It will undoubtedly be fundamental for this task to be able to plan and study the circumstances of each specific case, but it is important to take as reference the need to share and join forces in searching for the balance sought by all.

a

Ma´laga-Marrakech Declaration on Groundwater in the Mediterranean, 2006. (This Declaration is the result from two international congresses organized in 2006, AQUAinMED’06 – Ma´laga – and GIRE3D – Marrakech). b The United Nations Economic Commission for Europe Convention on access to information, public participation in decision making, and access to justice in environmental matters, Aarhus (Denmark), on 25 June 1998. c Johannesburg Declaration on Sustainable Development (World Summit on Sustainable Development, United Nations, 2002). d Commission on Sustainable Development, Report on the sixteenth session May 2007 and 2008 (Economic and Social Council, United Nations). Author: D. Sanz.

rational actions of individual actors, in our case groundwater users, lead to the demise of all, that is, aquifer over-use. This is because groundwater, which is a classic example of a common-pool resource, is defined by two characteristics: the

resource is largely rival and nonexcludable. These commonpool resources exist at different scales from transboundary to regional or small local aquifers. The works of Ostrom (1990) and other institutionalists have demonstrated that this case

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underestimated the capacity of the users to self-regulate their actions, that is, to develop rules in norm and rules in use to prevent aquifer overuse. The groundwater silent revolution described earlier in the chapter has however outpaced the capacity to develop institutions suitable for good groundwater governance in terms of resilience, while maintaining a level of flexibility and adaptability to cope with a high degree of change.

1.07.5.1 Groundwater Institutions: Mapping Groundwater Institutional Design A number of conditions have been well documented in the literature for the successful management of common-pool resources. These factors are summarized in Box 3 (Schlager and Lopez-Gunn, 2006) in relation to groundwater.

1.07.5.1.1 Boundary definition The first tenet of institutional theory refers to boundaries. This refers on the one hand to natural boundaries and on the other to institutional (property right) boundaries. In the first case, the definition of natural aquifer boundaries for management purposes has the added complication that groundwater aquifers do not necessarily coincide with surface-water systems. In addition, groundwater suffers from the same problem that surface water had traditionally experienced, lack of overlap between administrative and natural boundaries (i.e., problem of fit), that is, the boundaries of for example, regional administration do not coincide with river-basin boundaries, with the added twist that surface basins and aquifers often do not coincide, which further increases the complexity. An interesting example is currently pursued under the EU WFD, which has adopted a twin-track approach of managing water according to river basins while simultaneously mapping groundwater bodies, while setting the objective to achieve

good status for all water bodies in the EU (surface as well as groundwater) by 2015. According to Howe (2002), assigning well-defined groundwater property rights, for example, through pumping permits (discussed earlier) enhances the value of water, which creates incentives to use water more effectively or to transfer rights and/or use to third parties who are willing to pay for pumping rights. Therefore, the most complex challenge for water laws is the administration of water rights, that is, the granting of licenses, concessions, permits, and other legal deeds for the abstraction of groundwater, and for the discharge of waste water directly or indirectly into the aquifer. Groundwater, in particular, offers additional problems because of the following: first, the potentially large and often heterogeneous number of users, and second, the boom in use which has often overwhelmed administrations. In Spain, 20 years after the 1985 Water Law, the registration of groundwater rights was required (1988), but the administration has still not finished the process; or in places such as Mexico an ecological price has been effectively been paid for the process of registering 330 000 water rights by 2003, by over-allocating groundwater resources. The new 2002 Water Law in China established the need to obtain groundwater permits. Yet, the issuing of water permits in China by counties is proceeding very slowly, and there is lack of consistency between authorized abstractions via permits and groundwater-resource availability (Foster et al., 2004). Furthermore, growing experience in the process of assigning groundwater property rights has shown that it is crucial to take context into account when assigning groundwater rights, for example, in South Africa, where plural legislative frameworks (formal and customary) coexist. These plural, often dual legal systems have important implications for the registration of groundwater property rights, due to overlapping legal orders. The diversity and flexibility of customary laws, principles, and practices may be intentionally or

Box 3 Ostrom’s institutional design principles applied to groundwater institutions. Reproduced by Lopez-Gunn E from Ostrom E (1990) Governing the Commons: The Evolution of Self-Governing Irrigation Systems. Cambridge: Cambridge University Press; Schlager E and Lopez-Gunn E (2006) Collective systems for water management: Is the tragedy of the commons a myth? In: Rogers P, Llamas MR, and Martı´nez-Cortina L (eds.) Water Crisis: Myth or Reality?, pp. 43–60. London: Taylor and Francis; and Cleaver F and Franks T (2005) How Institutions Elude Design: River Basin Management and Sustainable Livelihoods. BCID research paper 12, ICID Conference, London. *

* *

* *

*

*

*

Clearly defined boundaries. Both the boundaries of the aquifer and the individuals or households with groundwater rights from the aquifer are clearly defined. This principle refers both to the physical boundary of the aquifer and a clear identification of groundwater rights (legal boundary on groundwater). Collective choice agreements. A clearly defined groundwater-user group or community should be involved in groundwater management. Appropriation rules. Operational rules in relation to time, location, technology, or groundwater-abstraction units should include the groundwater users affected by these rules and should be included in decision-making processes to modify these appropriation rules. Monitoring. Monitors who actively audit physical conditions and behavior are accountable to groundwater users and/or are groundwater users themselves. Graduated sanctions. Sanctions are devised for noncompliance with collective rules (operational rules). Groundwater users who violate operational rules are likely to receive graduated sanctions by other groundwater users, by officials accountable to these groundwater users, or by both. These sanctions have to be applied consistently, impersonally, and rapidly. Conflict-resolution mechanisms. Groundwater users and officials have access to low-cost local arenas to resolve conflict among users or between users and officials. Conflict-resolution mechanisms should be clear, accessible, and quick. Legitimacy. The legitimacy of groundwater users to organize and set up their own institutional arrangements is not challenged by external government authorities. Nested enterprises. Local groundwater institutions are nested within other levels of decision making, in multiple layers, which facilitate governance (in terms of consistent operational rules, monitoring, and enforcement and conflict resolution).

Groundwater Management

unintentionally replaced by new water laws, and uniform rigid principles and requirements. Legal frameworks empower if these recognize rights of existing water-user communities, and enable legal recourse if rights are harmed. Plural groundwater property-rights systems ideally have to be based on the principles of good governance: transparency, accountability, and the rule of law. There is also a risk in idealizing customary groundwater rights, which might not necessarily, for example, be gender neutral.

1.07.5.1.2 The role of groundwater-user associations Decentralization of groundwater-resource management is coherent with the creation of collective institutions like groundwater-user associations that can be directly involved in groundwater management. There is a range, nevertheless, on the degree of management devolution to groundwater-user groups, for example, market co-production, co-management, or regulated autonomy. In the late 1990s, decentralization was a consequence of rolling back the state, and transferring management directly to users – participatory-irrigation management (PIM; or irrigation-management transfer) since the dominant use of groundwater globally is agriculture. This is part of the wider trend in PIM (Merrey et al., 2007). In the case of groundwater, PIM has interesting twists and turns because at least two types of groundwater-irrigation systems can be identified: first, the case of collective wells which are managed as very small surface-water systems, and second, and most common, individual farmers exploiting their well for productive agriculture and/or livelihoods. The creation of wateruser groups and PIM would be similar to surface water in the first case, and would face similar limitations as those recently put forward for surface water, that is to say, that this is no panacea and it is suitable in some cases but not necessarily in all cases. These water-users associations (WUAs) are much smaller than WUAs for surface irrigation: this makes it simpler to organize them but it is also less important for them to be formal organizations. Informal groups are generally adequate for managing irrigation from individual wells, even when managed by groups of up to 50 farmers, such as in China. The second case is a true case of collective action because individual users have to be persuaded externally (top-down) or realize (internally) that the benefits of self-organization are higher than the costs, and that free riding on the collective action of others is now penalized either through formal sanctions or through informal, social norms. The objective in this case is regulation of the aquifer (i.e., the source of water) rather than equitable management of the distribution of water from the source. The case of PIM in groundwater is fascinating because there is evidence from groundwater-user associations that have been both created top-down and others which have emerged bottom-up, spontaneously. This is the case of Spain, where Comunidades de Aguas Subterraneas – which is in effect, part of the water authority and instigated by the administration, and which manages groundwater as part of the public domain – coexists with Comunidades de Usuarios de Aguas Privadas, where private groundwater-user groups have been created through user initiative. More research is needed on delivery of management outcomes; what is already evident is that the scale of

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these groundwater-user groups is large, managing aquifers which can cover areas from 7000 km2 to 300 km2, and where success is mixed in terms of sustainable aquifer management. In China, groundwater management has gradually become more decentralized, and bottom up, with increased stakeholder participation at all levels, and closer interaction between users at the local level and the responsible authority, the Water Resource Bureau, normally designated at county level for rural areas and district level for urban municipalities, while some groundwater-user associations have also been established (Foster et al., 2004). The few examples of groundwater-user associations that have become effective resource managers have two things in common: they have successfully articulated common goals and objectives, and they have established mutually accepted rules regarding resource access and use, in order to guarantee the long-term availability of groundwater to users. For example, in Mexico, in the early 1990s, due to intensive groundwater use in the central and northern part, many groups started to emerge concerned with the problem of intensive groundwater use and negative externalities: for example, the spontaneous creation of the Grupo del Agua in the Comarca Lagunera (1991) and the Grupo del Agua of Santo Domingo valley a year later (1992). Other groups appeared in other areas. Initially, there was lack of clarity on the regulatory structure of these groups and their financing, which meant there was little support from the federal level. Initially, the Mexican Federal Government did not legitimize these spontaneous water-user groups until the mid-1990s, when these groups reorganized themselves as Comites Tecnicos de Aguas Subterraneas or COTAS, starting in the Queretaro valley, and then spreading to other aquifers in the central and northern part of Mexico. In the state of Guanajuato, local authorities encouraged the formation of COTAS in all aquifers in the state, supporting them financially (Escolero and Martinez, 2007). Meanwhile in the USA, local landowner associations in Texas have been experimenting with the feasibility of selfmonitoring and regulation under local groundwater districts, which would set pumping limits and well placement based on hydrologic models, to deliver public goods such as open-space protection and aquifer recharge through cooperative landowners associations (Wagner, 2005). In India there is evidence of the spontaneous creation of WUAs, through what Shah (2005) calls swayambhoo (self-creating), involving entrepreneurial efforts, which are normally present since most groundwater users are by definition smallscale entrepreneurs. It is estimated that over a quarter of Indian irrigated areas operate through this kind of spontaneous creation of informal water markets (Shah, 2005). The challenge is when swayambhoo institutions have to be scaled up, whether motivation can shift to longer term, and collective self-interest, then, can also start to internalize externalities (Box 4).

1.07.5.2 An Institutional Audit of Groundwater Institutions Added research and experience have however highlighted new dimensions to the institutional framework analyzed above (Cleaver and Franks, 2005; Cleaver and Franks, 2008). The socalled post-institutionalist turn has added some caveats and

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Box 4 WUAs for groundwater management in China. By Simon Howarth (UK DFID-funded Water Resources Demand Management Assistance Project. Water-users associations (WUAs) are being set up in each village in the Shiyang River Basin, an arid internal river basin in Gansu Province of Northwestern China. They are being promoted by the government to assist in the management of groundwater, but with a primary focus on achieving water savings. Existing tube-well management arrangements, by local production groups (subdivisions of villages) supported by water-management stations (WMSs) (government) at township level, are believed to be effective and equitable, but insufficiently focused on reducing total water use – with the result that groundwater levels are dropping at 50– 100 cm yr1, making agriculture unsustainable. These WUAs thus have different objectives to WUAs set up elsewhere in the world, which are required to improve management of large surface canal flows and hence ensure greater equity of water distribution. Each village (or WUA) typically includes 20–50 tube-wells (each serving 5–20 ha, farmed by 10–50 households) which are managed by production groups (water-user groups). The tasks of the village-level WUA include assistance to the WMS in many of the new groundwater-management procedures, such as issuance of household water-rights certificates, enforcement of permits, and collection of fees – all of which are aimed at reducing the amount of water that farmers use. These are onerous requirements and thus the WUA are repaid part of the water-resource fees collected in order to cover a small salary for directors and vicedirectors and some administrative costs – in recognition of the role that WUAs play in water-resources management. This formal process of paying staff from part of the newly introduced water-resources fee is important for ensuring that the WUAs are effective and sustainable. In addition to these responsibilities for assisting the government, WUAs also have a small role in water management which includes improving maintenance; reducing conflicts; planning, implementing, and monitoring water distribution; monitoring groundwater levels; and ensuring effective communication between WMS, WUA, and farmers. Much of this work is done by well-established informal means by production groups, but the WUA coordinates between production groups and provides services at a higher level – such as employing a maintenance technician who is available to all groups, linking groups to governmentsponsored training programs, and assisting in contracts with crop-grower associations, seed suppliers, and markets. These WUAs are intended to be independent, autonomous, democratic, village-level organizations, but for practical reasons, the staff are often largely drawn from existing village committees (these are elected, but all candidates are required to be vetted and approved in advance). On paper, it is a strong system, but it is newly established and not yet fully effective. Many questions remain unconfirmed, including sustainability of financial arrangements, ability to deliver a positive service to farmers, and the willingness of farmers to accept the restrictions on water use. Further work on WUAs will require a combination of administrative measures at provincial, municipal, and county levels, and capacity building among WUAs. This capacity building will in turn require awareness-raising at the various levels of government, where there is typically greater faith in top-down controls (such as IC cards) or infrastructural improvements (canal lining) than there is in local institutional methods for water savings. Nevertheless, early indications are that the strong commitment to water savings by the government will ensure that WUAs will be effective, but that their role and responsibilities will be modified and simplified as they are implemented.

new dimensions to a strict application of Ostrom’s institutionalist framework. The main criticisms are that it is does not provide a causal analysis for the processes underlying these design principles. In particular, the areas that are increasingly perceived as fundamental to sound groundwater governance are: first, the key role of social capital and higher-level authorities; and second, the relevant role of political leadership and acknowledging the politics and vested interests of groundwater use, which are played out in the prioritization of groundwater use among competing users; and third, the potential problem of corruption as a symptom of a malfunctioning groundwater systems and the antidote of transparency and participatory groundwater management.

1.07.5.2.1 Higher-level authorities: Supporting, legitimizing, and leading The relevance of higher-level authorities comes to the fore as an essential supporting element for effective institutions and the development of organizational capacity since both authority structures and social norms (e.g., collective action by users) have to support and underpin the functioning of design principles. Higher-level authorities are key as facilitators for local groundwater management and for the vertical integration between the different spatial scales (farm level, aquifer, and regions, national, and international scales). For example, it appears that cross-scale linkages exist in China where there is a provision for transboundary issues across provinces, and also indirect leadership, as professional guidance from higher-level

authorities without any hierarchal subordination. County government can issue groundwater regulations within its boundaries, in agreement with provincial and national legislation (Foster et al., 2004). Higher-level authorities are increasingly perceived as a necessary condition to support local institutional arrangements. One of the most important roles for higher-level authorities is either to provide leadership or to facilitate leadership. In many cases, leadership is actually in the form of legitimizing or supporting local leaders. These local leaders in turn can drastically reduce the transaction costs of institutional change. In India, for example, the common aspect of all successful tank institutions was a leader or a leadership compact, which could sway the community and thus drastically reduce the transaction costs of ‘‘enforcing institutional arrangement that would either not work in their absence nor survive them’’ (Shah, 2005, p. 17).

1.07.5.2.2 Transparency and participatory groundwater management This chapter starts from the tenet that there is no global physical groundwater crisis; rather, there is a crisis of groundwater governance. Governance in this context is defined as the interplay of actors (public, private, and civic) to promote societal goals and the production of collective goods. One of the key basic assumptions of effective functional groundwater management is transparency and participation by all groundwater users in the decision-making process in

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line with the Aarhus Convention on Public Participation (UNECE, 1998). The problem of corruption often damages those most vulnerable, weakening the rule of law, and fostering social norms that systematically prioritize private gain over social well-being. Corruption according to Transparency International is about breaking socially established expectations of appropriate behavior (Stalgren, 2006). As stated earlier, groundwater has some inherent characteristics that should make it less prone to corruption. In the case of groundwater, the timescale and size of investment is normally smaller than in the case of surface-water projects. Evidence of corruption in the case of groundwater tends to refer to drilling concessions, bribing meter readers, distorted site selection for boreholes, for example, for those with more political or economic influence; bribery to obtain drilling permits or to cover up excessive abstraction, to obtain preferential treatment for services or repairs, and also to falsify meter readings (Transparency International, 2008). Advances are constantly made to facilitate transparency, accountability, and decentralization in groundwater management and use, for example, in the use of technology (Calera et al., 1999). Three measures are considered crucial in the case of effective groundwater management. First, reduce the complexity in regulation, licensing, and control; this is to prevent a weak and ineffective legal system which can encourage a clientelistic patronage system. Second, facilitate and incentivize so-called participatory monitoring. As stated early, transparency, monitoring, and sanctioning are part of healthy groundwater institutional arrangements. Robust groundwater institutions can benefit from advances in participatory geographical information systems, that is, use of technology jointly by groundwater users and regulators to increase transparency in water use and allocation. A good example is currently being implemented in the Mancha region in Spain, where satellite information is being used directly by farmers through an irrigation advisory service, which integrates realtime data to help farmers improve water use by different crops, while optimizing production (Calera et al., 1999). Third, encourage transparent access to data on groundwater use, licensing, and subsidies. This can be strengthened by partial decentralization to water users, to involve them in decision making, which would decrease the transaction costs of obtaining good-quality information while increasing the level of information available (Box 5). In summary, good, symmetric information equally accessible to both users and the regulator is crucial to facilitate cooperation among aquifer stakeholders. This information

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ideally has to be externally audited and contrasted, to allow for advocacy and disclosure of illicit behavior. Often, easy access to good and reliable data on abstractions, water quality, and aquifer water levels is a prerequisite to succeed in groundwater management. Current information technology can help information to be made easily and economically available to an unlimited number of users. Nevertheless, in a good number of countries, it will be necessary to change the traditional attitude of water agencies of not facilitating easy access to water data to the general public. This partly comes from a shift in mentality that strengthens accountability of public authorities downward to users and civil society.

1.07.6 The Complex Concept of Groundwater Sustainability and Future Management Issues An economic, efficient use of an aquifer would imply maximizing the present value of the resource in the case of the Ogallala, in Texas, where abstractions are much greater than natural recharge. It was discussed earlier in Section 1.07.2.5 that this exemplifies the complexities in defining what is meant by sustainable groundwater management. The exhaustible nature of the resource would raise the issue of appropriate long-term economic and demographic development of the region. The availability of open-access inexhaustible resources such as groundwater often invites gold-rush patterns of excessive fast exploitation and maladapted patterns of infrastructure and social development, that is, so-called boom towns (Howe, 2002). Economic efficiency in a renewable aquifer may imply drawing down the aquifer during droughts and allowing its recharge in periods of good surface flow. However, ecological dimension of sustainability used to equate recharge equal to abstractions is what some authors consider the renewable yield. Nevertheless, the EU WFD introduces a more complex concept, the achievement of good ecological health of aquatic ecosystems, which depends on the available yield. This new concept, not fully applied yet, may imply significantly lower amount of groundwater allowed abstraction than the renewable yield. Literature, such as Moench (2003) and Shah et al. (2007), and examples in Indonesia, China, USA (Ogallala), Spain, and Mexico, seem to indicate that de facto development (e.g., in terms of agricultural productivity) is prioritized over longerterm ecological groundwater sustainability. In Indonesia, regulation of groundwater is perceived to be at cross-purposes with industrial growth. In China, pure economic growth is

Box 5 Participatory groundwater monitoring in China. By Simon Howarth, based on Gansu under the UK DFID-funded Water Resources Demand Management Assistance Project. Simple monitoring by villages of the volumes of water abstracted and of the groundwater level is valuable for developing an understanding of groundwater. This has three classes of benefit: promoting awareness of groundwater, which is commonly less well understood than surface water; this can profitably be incorporated into school curricula so that children become aware of water issues: * * *

Enabling WUAs to manage their resource better, and understand why restrictions are being introduced. Providing data to supplement formal data-collection programs by government hydrology bureaus. Involving communities increases ownership of the concepts and reduces asymmetry of information. The information can easily be published on village notice boards.

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seen as a central part of development policy, in the policymakers’ mindset toward pure economic accumulation. In Mexico, continued overdraft in Hermosillo is driven by problems of reconciling economic efficiency and ecological sustainability (e.g., problem of saltwater intrusion), and where huge economic returns derived from groundwater in terms of income create an incentive to search for additional water resources through surface-water transfer (Escolero and Martinez, 2007). This is similar to the case in Southeast of Spain, in the Murcia and Almeria regions, where the intensive use of aquifers for highly productive agriculture started to drive national water policy to transfer water to this area, because of the political difficulty of controlling this intensive but highly profitable intensive groundwater use (Llamas and Martı´nezSantos, 2005; Llamas and Martı´nez-Cortina, 2009). The strength of negative externalities depends partly on aquifer characteristics (e.g., transmissivity storage), spacing of wells, and connection to surface water. These questions over preferred criteria as against competing uses go to the heart of the meaning of groundwater sustainability and to what extent this is feasible or, indeed, it is perceived as feasible or possible due to the institutional path dependencies of choices made in the past.

1.07.6.1 Groundwater Management Externalities In groundwater, externalities are the rule rather than the exception. The real issue is not the elimination of externalities (usually physically impossible) but rather, whether the impacts on third parties are excessive according to certain criteria. The relevant policy question is whether these externalities are considered excessive and for which criteria they are used or prioritized: economic efficiency, groundwater sustainability, or Table 5

social equity. In the EU, according to the WFD, the goal is to restore the ecosystems to a good ecological status unless the cost of this recovery is economically or socially very difficult. In this case, member states have to report in detail to the European Commission on the extenuating circumstance to ask for derogations. This is a process currently underway, but it appears that countries are anticipating the difficulty of complying with the WFD by 2015. Here, we summarize five indicators of typical problems of intensively used aquifers, but it is important to mention that these are sometimes used inadequately (Table 5).

1.07.6.1.1 Degradation of groundwater quality Groundwater abstraction can cause, directly or indirectly, changes in groundwater quality. The intrusion into a freshwater aquifer of low-quality surface water or groundwater, because of the change in the hydraulic gradient due to groundwater abstraction, is a frequent cause of quality degradation. This degradation of groundwater quality may not be related to excessive abstraction of groundwater in relation to average natural recharge. Other causes may be responsible, such as return flows from surface-water irrigation, leakage from urban sewers, infiltration ponds for wastewater, septic tanks, urban solid-waste landfills, abandoned wells, mine tailings, and many other activities not related to groundwater development (Custodio, 2002). For instance, the groundwater-quality degradation in many Central and Northern European countries is related to intensive rainfed agriculture. Saline intrusion may be an important concern for the development of aquifers adjacent to saline water bodies. This is a typical problem in many coastal regions of semiarid or arid areas. Moreover, in this case, the relevance of saline-water

Typology of groundwater externalities

Type of externality

Externality

Explanation

Environmental

Affected ecosystems

Socioeconomic

Pumping costs externality

Damage to ecosystems or surface-water features dependent on discharge from aquifers; spring-flow reduction Increase in pumping costs due to drop in aquifer levels, these costs can be fixed or marginal costs that one user imposes on another when pumping lower from a water level, external costs can be reduced by selecting a better well location Due to salinization or marine intrusion Water quality varies with depth (normally more saline with more depth) and also location specific, for example, aquifer located in coastal areas and islands (e.g., saltwater intrusion), the spread of low-quality water within an aquifer Water pollution due to intensive agriculture, for example, with nitrates and/or pesticides Decrease in pore-water pressure, related to amount of groundwater withdrawn Aquifer compaction with possible resultant reduction in aquifer storage capacity For example, when small farmers cannot adjust to the drop in aquifer levels Increased opportunity cost due to an increase in scarcity value due to intensive use Reduction to buffer value provided by groundwater against drought Diminished economic activity in the area, reduced water availability for other water-right holders, and reduced land-use options for future inhabitants

Potential loss of agricultural land Groundwater quality externalities

Land subsidence Aquifer compaction

Option and intertemporal externalities

Social externalities Increased scarcity value Buffer value of a groundwater stock Intertemporal externalities

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intrusion depends not only on the amount of the abstraction in relation to the natural groundwater recharge, but also on well-field location and design, and the geometry and hydrogeological parameters of the pumped aquifer. In most cases, the existing problems are due to uncontrolled and unplanned groundwater development and not to excessive pumping. It appears, for example, that the last half-century seawater intrusion has been well controlled in the coastal plains of Orange County (California) and Israel.

1.07.6.1.2 Susceptibility to subsidence When an aquifer is pumped, the water-pore pressure decreases and the aquifer solid matrix undergoes a greater mechanical stress. This greater stress may produce compaction of the existing fine-grained sediments (aquitards) if the stress due to the decrease in water-pore pressure is greater than the so-called preconsolidation stress. This situation has occurred in some aquifers formed by young sediments, such as those in Mexico City, Venice, and others. In Bangkok (Thailand), parts of the city were sinking at a rate of 10 cm yr1, with an increased risk of flooding and damage to roads and buildings. Caves and other types of empty spaces may exist under the water table in karstic aquifers. When the water table is naturally depleted, the mechanical stability of the roof of such empty spaces may be lost and the roof of the cave collapses. This is a natural process that gives rise to the classical dolines and poljes in karstic landscapes. When the water table depletion or oscillation increases due to groundwater abstraction, the frequency of karstic collapses can also increase. There are a number of well-known examples of land subsidence due to intensive groundwater use. In both cases, the amount of subsidence or the probability of collapses is related to the decrease in pore-water pressure, which is related to the amount of groundwater withdrawal. Nevertheless, the influence of other geotechnical factors may be more relevant than the amount of water abstracted in relation to the renewable groundwater resources of the aquifer. In Tianjin city (China), excessive abstraction of deep groundwater caused a land subsidence of up to 3.0 m (Foster et al., 2004). In Texas, Galveston and adjacent counties have experienced subsidence due to the long-term drop in aquifer levels (Wagner, 2005). Meanwhile, land surfaces in parts of central Arizona have fallen by 20 m in the last 20 years (Howe, 2002).

1.07.6.1.3 Interference with surface water and ecological impacts There are potential conflicts due to groundwater pumping and its interaction with surface waters and riparian habitat. For example, if there is no source of capture, a well will continue to withdraw water from storage until, either the aquifer is depleted, or the drawdown exceeds the well depth. The groundwater pumped from an aquifer is derived from a decrease in storage in the aquifer, a reduction in previous discharge from the aquifer, an increase in the recharge, or a combination of these changes. Capture may be defined as the increase in recharge plus the decrease in discharge. Examples of capture are: (1) an increase in groundwater recharge from losing streams (or increased infiltration); (2) a decrease in groundwater discharge to gaining streams (or interception of

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baseflow); and (3) the reduction in the component of evapotranspiration that is derived from the saturated zone. If we restrict the groundwater pumping to the capture, there will no longer be a decrease in storage. Restricting the groundwater withdrawals to what may be captured is a definition of safe yield or sustainable yield for the aquifer. Nevertheless, the definition of safe yield is controversial for some authors. Groundwater mining, that is, when water is abstracted mainly from storage (as discussed in Sections 1.07.2.5 and 1.07.2.6) may also be considered a safe yield, which can be valid under certain conditions. This is because sustainability has to consider different aspects, including economic, ecological, and social, and the real-life difficulties of implementing the concept of sustainability. Most river systems have a hydrology that is simple in concept but complex in detail. Some anthropogenic activities may have a significant impact on the catchment hydrologic cycle. For instance, the intensive use of groundwater for irrigation in the Upper Guadiana basin (Spain) has resulted in serious water-table depletion (B30–40 m). The most alarming consequences of the water-level drop were changes in the groundwater flow patterns and in the form, function, and quality of many wetlands. Areas that had received the natural discharge from the aquifer became natural recharge zones (Herna´ndez-Mora et al., 2003). This has produced a spectacular decrease in total evapotranspiration from the water table and wetlands, evaluated between 100 and 200 Mm3 yr1 (Martı´nez-Cortina, 2001). From the point of view of the water budget, there is an important increase (almost 50%) of the annual renewable resources, understood as the water that can be abstracted from the aquifer maintaining the water level as in the previous year, and calculated as the difference between aquifer recharge from precipitation and losses from evapotranspiration. This artificial depletion of the water table can also change dramatically aquifer–streams relationship, as in the previous example. Gaining rivers fed by aquifers may become dry except during storms or humid periods when they may become losing rivers, an important source of recharge to the aquifer. Nevertheless, this new water budget may present legal problems if the downstream water users have previous water rights (Llamas and Martı´nez-Cortina, 2009). The ecological impacts, mainly caused by water-table depletion as it has been showed in the Upper Guadiana basin case, are becoming an important new constraint in groundwater development in some countries, especially in the 27 countries of the EU because of the requirements of the WFD. A famous case is the Tablas de Daimiel National Park, a Ramsar site, whose main source of water used to be aquifer discharge before intensive irrigation made the area a recharge area rather than a natural outflow for the aquifer (Figures 5 and 6). Decreasing or drying up of springs and wetlands, low flow of streams, disappearance of riparian vegetation because of decreased soil moisture, alteration of natural hydraulic river regimes, and changes in microclimates because of the decrease in evapotranspiration, can all be used as indicators of ecological impact. Reliable data on the ecological consequences of these changes are not always available, and the social perception of such impacts varies in response to the cultural and economic situation of each region. The lack of adequate

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Figure 5 The Tablas de Daimiel National Park, October 2008. Photo courtesy: Pedro Zorrilla).

Figure 6 The Tablas de Daimiel National Park, October 2008. (Photo courtesy: Pedro Zorrilla).

scientific data to evaluate the impacts of groundwater abstraction on the hydrologic regime of surface water bodies makes the design of adequate restoration plans difficult. For instance, wetland-restoration programs often ignore the need to simulate the natural hydrologic regime of the wetlands, that is, restore not only its form but also its hydrological functions (Bergkamp and Cross, 2007). Similar problems result in trying to restore minimum low flows to rivers and streams. Oftentimes, minimum stream flows are determined as a percentage of average flows, without emulating natural seasonal

and year-to-year fluctuations to which native organisms are adapted (Llamas and Garrido, 2007; Garrido and Llamas, 2009).

1.07.6.2 Groundwater: Future Risks and Opportunities for Management 1.07.6.2.1 Groundwater and climate change The latest report of the IPCC by Bates et al. (2008) only marginally addresses the role groundwater can play in the

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adaptation or mitigation of the potential negative effects of climate change. The report says ‘‘There is a need to improve understanding and modeling of climate changes related to the hydrological cycle at scales relevant to decision making. Information about the water related impacts of climate change is inadequate – especially with respect to water quality, aquatic ecosystems and groundwater (added emphasis) – including their socio-economic dimensions’’ (p. 4). A detailed analysis of this topic is outside of the scope of this chapter. However, it seems that the role of groundwater development will increase significantly if the pessimistic predictions of the IPCC reports for the arid and semiarid areas become true. These pessimistic forecasts are mainly related to the increase in evaporation rates, due to the increase in temperature, and to the higher drought frequency. As it is shown in this chapter, the evaporation of groundwater from the aquifers is usually irrelevant and one important property of most groundwater reservoirs or aquifers is their resilience to dry spells. The UK Groundwater Forum, for example, studied potential scenarios for groundwater as a result of climate change, and some of these scenarios pointed to a long-term decline in aquifer storage, increased frequency, and severity of groundwater-related floods, mobilization of pollutants due to seasonally high water tables, and saline intrusion due to sealevel rise (Bergkamp and Cross, 2007). However, these predictions have to be contrasted with others across the world, with different climatic regimes. The Edwards aquifer is one of the largest freshwater aquifers in the USA with a total area of 15 640 km2, and a primary source of water (agricultural and municipal) for southern Texas (Loaiciga, 2003). It has been identified as one of the areas most vulnerable to complex, nonlinear climate feedbacks, and where potentially aquifer-exploitation strategies must be adapted to climate variability. In fact, when climate and groundwater-use changes are considered together, the role of groundwater use over climate prevails, that is, changes in groundwater use due to population growth and changes in land use or economic preferences may cause more profound aquifer impacts than those associated with global warming. For example, climate change in San Marcos springs could increase spring flow relative to the base condition by 17%, while the groundwater use alone in the year 2050 can reduce springwater flow by 22%, that is, groundwater use dominates over climate change. Therefore, the primary threat to the Edwards aquifer comes from the rise in groundwater use associated with predicted growth not from climate change. The latter in fact would increase spring flow in the study area (Loaiciga, 2003). There is also increased understanding that the vegetation response to climate change could either increase or decrease recharge. Climate change in fact could increase aquifer recharge according to recent simulation models, although highly dependent on geological settings (Green et al., 2007). In study areas characterized by sandy top soils and large interconnected aquifers, groundwater levels rose significantly. In Australia, simulations of twice the existing CO2 led to significant changes in the rate of groundwater recharge in Mediterranean and subtropical climates. Water recharged from 34% slower to 119% faster in the Mediterranean climate and from 74% to 500% faster in the subtropical climate.

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Opportunity for decrease exists but the general trend is toward increase in recharge (Green et al., 2007). In the context of groundwater and climate change, it is important to note the spatial and timescales of aquifer and climate systems. First, aquifers often operate in scales of much less than 106 km2 and in a great majority of cases, groundwater basin encloses areas of 104 km2, while global climate models (GCMs) operate on 200 km  200 km (4.104 km2) and regional climate models (RCMs) with resolutions of the order of 20 km  20 km (Loaiciga, 2003); therefore, the scales do not necessarily match up. Second, the nature of medium- and long-term climate predictions and the contrast between floodimpact studies with temporal scale from minutes to days, and drought impact studies precipitation and temperature temporal scale from days to years depending on inter- and intraseasonal variation. These uncertainties in both space and time make predictions, in terms of climate change and variability, difficult; what is clear is that in this context, aquifers have the natural capacity to act as climate regulators, that is, buffering capacity for drought and floods.

1.07.6.2.2 Future management issues There are a number of future management issues that become apparent and whose importance is increasing. One of the main issues is the joint use of surface and groundwater and the linkages between water quality and water quantity. There are good examples across the world of successful joint management of surface and groundwater that play to the strengths and weaknesses of both. For example, the case of Israel, and the case of the cities such as Barcelona (Spain) and Phoenix (Arizona), which rely on groundwater supplies as a strategic resource in times of drought. However, there are also examples on lack of joint management or in fact disjointed management, that is, when poor groundwater management leads to surface-water transfers to compensate. This is the case, for example, of lobbies pushing for surface-water transfers or the authorities pushed to find additional water supplies due to groundwater-quality problems. For example, both in London and in the coastal metropolitan area in Barcelona, there is a problem with aquifer rebound, and these polluted groundwater resources signify that it is easier to invest on large desalination plants to augment supply, since the costs of cleaning polluted groundwater are prohibitive. Meanwhile, in both Spain and Mexico, pressure builds in areas where aquifers are intensively used to bring surface-water supplies. The city of Hermosillo (Mexico) is heavily dependent on groundwater for its level of productivity and the residents of the area have been lobbying the Mexican government for a large water transfer that would bring water from the State of Sinaloa 485 km away (Plan Hidra´ulico del Noroeste). As groundwater storage continued to fall, the rising shadow (marginal price) value for groundwater rose. This shadow value can indicate what would be the optimum timing for the water transfer to occur, which would be when the values of the transferred water are lower (US$ 0.0222 m3) than the shadow groundwater value price (US$ 0.0224 m3). This highlights that the ideal time to build the project would be in the 29th year. This indicates that if the water project was built at the current costs of groundwater abstraction, it would

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have been built prematurely (Howe, 2002). This, however, is only if one applies economic-efficiency criteria. If other criteria (such as environmental sustainability) were used, this would call for a much quicker solution (e.g., due to the external costs of saltwater intrusion). In Spain, lack of groundwater management in the southern Mediterranean coastal belt, triggered in large, the conflict to divert water from the Ebro river in the north to help compensate for rapidly depleting aquifers (Llamas and Martı´nez Santos, 2005). Water agencies tend to build projects far in advance of their justifiable need on pure economic terms (Howe, 2002), and often fail to capitalize on the synergy between effective joint surface and groundwater use, which by definition implies the management of both. Restrictions on groundwater allocations are a direct loss attributed to decision makers and thus unpopular, whereas a loss of income due to over-abstraction is a probabilistic loss. Therefore, it is politically rational for decision makers to prefer users to continue pumping than to take the (unpopular) decision to cut allocations and instead opt for the politically more popular water transfers. There are very few systems of explicit conjunctive management. Until the 1940s, one main reason was the lack of understanding of hydrogeological knowledge and therefore the poorly developed model of surface and groundwater interaction. There are recent examples of regulatory innovations to deal with groundwater due to a much clearer understanding and a capacity to play to the strengths of surface and groundwater joint use – for example, in Colorado, where tributatory groundwater rights have been incorporated into the prior appropriation (priority) doctrine of water law, which in theory could preclude conjunctive management since groundwater tends to be junior rights compared to surface senior water rights (Howe, 2002).

1.07.6.2.3 Groundwater: Issues of fit and political windows of opportunity There is increasing evidence of the institutional diversity of groundwater user groups, from landowners associations, local landowner cooperatives, natural resource cooperatives, and water districts in Texas, to tube-well cooperatives in India, to agricultural transformation societies, and both public and private user communities in Spain. There is also some evidence that spikes in groundwater scarcity can trigger organization of these groundwater-user groups and in fact provide a window of opportunity for collective action, for example, drought can act as a motivator for self-organization or increased competition for groundwater resources among users. Drought intensifies conflicts and yet stimulates short-term and long-term efforts to modify rules and procedures for regulating rights. Scarcity value encourages the spontaneous creation of cooperative groups as groundwater becomes scarcer and groundwater economic value raises the cooperative model of groundwater pumping. Increasingly, it is appreciated that issues such as droughts rather than perceived as short-term crises are in fact also an opportunity for institutional change and adaptation. These are windows of opportunity for political action and social change since most stakeholders are receptive to the need for effective responses – that is, they provide opportunities for institutional innovation.

Strategically, choices can be made on the value of groundwater resources and prioritization of use, for example, for domestic water supply while at the same time providing incentives for economic transition. In addition, there is increased recognition on the importance of context and that there are no ideal aquifer-management regimes; rather governance arrangement. In particular groundwater basins are highly individualized, there is no single best-practice model for groundwater management; rather there are context-specific, multiple management scenarios which have to be negotiated. Institutional solutions that are viable in a particular context have to be framed within the inherent limitations of scientific knowledge, and centered on core objectives rather than specific groundwater parameters (Moench, 2007). That is, focus on livelihoods and environmental values rather than sustainable yield, which by itself is increasingly a contested concept (Llamas and Garrido, 2007). Responses have to be suitable for specific socioecological context rather than politically correct integrated management, which sometimes can be too rigid or overtly focused on technical ideals. Management in groundwater has to be pragmatic because timescales in the case of groundwater vary, on the one hand between the resource itself, which has an inbuilt lag, and requires a long-term perspective, and groundwater users themselves who often operate on a much shorter timeframe, normally driven by economic development. These two (long and short-term timescales) somehow have to be synchronized. The stabilization of the North China plain aquifer will be a long-term process when one considers that in 1988 the Hai river basin exceeded recharge by some 8800 Mm3 yr1, and an average recharge deficit of 40–90 mm yr1 (Foster et al., 2004). As discussed earlier in the chapter, at times it might be rational to use an aquifer intensively due to the associated socioeconomic and generational changes, which in time might decouple livelihoods from groundwater dependence, for example, through education or a more diversified economy. Issues of spatial scale and fit as raised earlier are relevant since often human institutions do not match groundwater boundaries. In the case of groundwater, in many cases, it is the individual, micro-level which drives groundwater use as the aggregate demand of thousands of groundwater users. To this, one has to add the high levels of uncertainty, which are an integral part of groundwater management, possibly magnified due to climate variability and change. For groundwater institutions, these limitations are inherent rather than situational (Moench, 2007).

1.07.7 Conclusion Suggestions or recommendations to achieve sustainable and ethical groundwater management have been presented in many conferences. The Alicante Declaration (Ragone and Llamas, 2006) is one of the recent ones. We include here what we consider relevant aspects: 1. There is no doubt that agriculture is the main blue and green water consumer. The virtual water-trade analysis

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2.

3.

4.

5.

6.

7.

seems to show that in many countries the motto ‘more crops and jobs per drop’ is changing to ‘more cash and nature per drop’. Usually hydrological and economic productivity of groundwater irrigation is significantly greater than the corresponding productivities of surface-water irrigation. Detailed analyses on this aspect in different climates and countries should be done to confirm or reject these preliminary data. The spectacular increase in groundwater use that has occurred in the last five or six decades can be classified as a silent revolution because it has taken place with scarce planning and control by governmental agencies. This is the main cause of some observed negative impacts, which could be avoided or mitigated with adequate groundwater management. It is extremely difficult to provide a general guide to good groundwater management, as complying with all the different dimensions may not be possible in most cases. Emphasis on one or another is likely to depend on economic, social, cultural, and political constraints. Groundwater management requires a higher degree of user involvement than surface-water developments. Experience shows that sustainable aquifer use cannot be solely achieved by means of top-down control-and-command measures. User participation requires a degree of hydrogeological education which is still absent in most places. Steps should be taken to make the peculiarities of groundwater resources known to all, from politicians and water decision makers to direct users and the general public. This should begin at the school level. Appropriate groundwater management requires a significant degree of trust among stakeholders. This implies that groundwater data should be transparent and widely available (e.g., via the Internet). In addition, the system should be able to punish those who act against the general interest.

References Acharya G and Barbier EB (2000) Valuing groundwater recharge through the agricultural production in the Hadejia-Nguru wetlands in northern Nigeria. Agricultural Economics 22: 247--259. Aldaya MM, Llamas MR, Varela-Oretga C, Novo P, and Rodriguez- R (2009) Challenging the conventional paradigm of water scarcity through the water footprint: The Spanish example. In: Hoekstra A (ed.) Global Water Governance. London: Earthscan. Bates BC, Kundzewicz ZW, Wu S, and Palutikof JP (2008) Climate Change and Water, Technical Paper of the Intergovernmental Panel on Climate Change, 210pp. IPCC Secretariat, Geneva. Bergkamp G and Cross K (2007) Groundwater and ecosystem services: Options for their sustainable use. In: Ragone S, de la Hera A, Hernandez-Mora N, Bergkamp G, and McKay J (eds.) Global Importance of Groundwater in the 21st Century: The International Symposium in Groundwater Sustainability, pp. 233–246. Alicante, Spain, 24–27 January 2006. Westerville, OH: National Groundwater Association Press. Briscoe J (2005) India’s Water Economy: Bracing for a Turbulent Future. Washington, DC: World Bank. Brown GM (2000) Renewable natural resource management and use without markets. Journal of Economic Literature 38(4): 876--915.

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Llamas MR and Garrido A (2007) Lessons from intensive groundwater use in Spain: Economic and social benefits and conflicts. In: Giordano M and Villholth KG (eds.) The Agricultural Groundwater Revolution: Opportunities and Threat to Development, pp. 266--295. Wallingford: CABI. Llamas MR and Martı´nez-Cortina L (2009) Specific aspects of groundwater use in water ethics. In: Llamas MR, Martı´nez-Cortina L, and Mukherji A. (eds.), pp.187– 204. Leiden: CRC Press/Balkema. (ISBN 978-0-415-47303-3). Llamas MR, Martı´nez-Santos P, and de la Hera A (2007) Dimensions of sustainability in regard to groundwater resources development: An overview. In: Ragone S, de la Hera A, Hernandez-Mora N, Bergkamp G, and McKay J (eds.) Global Importance of Groundwater in the 21st Century: The International Symposium in Groundwater Sustainability. Alicante, Spain, 24–27 January 2006. Westerville, OH: National Groundwater Association Press. Llamas MR, Martı´nez-Santos P, and de la Hera A (2008) Hydropolitics and hydroeconomics of shared groundwater resources: Experience in arid and semiarid regions. Paper presented in the conference of the NATO Advanced Study Workshop. Varna, Bulgaria, 1–12 October 2006. In: Darnault C (ed.) Overexploitation and Contamination of Shared Groundwater Resources, pp. 415– 431. Dordrecht: Springer. Llamas MR, Shah T, and Mukherji A (2006) Guest Editors’ preface. Hydrogeology Journal 14(3): 269--274. Loaiciga H (2003) Climate change and groundwater. Annals of the Association of American Geographers 93(1): 30--41. Lopez-Gunn E (2003) The role of collective action in water governance, a comparative study of groundwater user association in La Mancha aquifers, Spain. Hydrogeology Journal 28(3): 367--378. Lopez-Gunn E (2009) Governing shared groundwater: The controversy over private regulation. Geographical Journal, 175(1): 39–51. doi: 10.1111/j.14754959.2008.00313.x. Lopez-Gunn E and Llamas MR (2008) Re-thinking water scarcity: Can science and technology solve the global water crisis? Natural Resources Forum 32: 228--238. Margat J (2008) Les Eaux Souterraines dans le Monde. 178pp. Orleans: BRGM e´ditions. Martı´nez-Cortina L (2001) Estimacio´n de la recarga en grandes cuencas sedimentarias mediante modelos nume´ricos de flujo subterra´neo. Aplicacio´n a la cuenca alta del Guadiana (Recharge Estimation in Large Sedimentary Basins Using Groundwater Flow Models. The Case of the Upper Guadiana Basin), 418pp. PhD Thesis, University of Cantabria, Spain. Melgarejo J and Molina A (eds.) (2005) Los mercados del agua. Ana´lisis jurı´dicos y econo´micos de los contratos de cesio´n y bancos de agua. Madrid: Civitas. Merrey D, Meinzen-Dick R, Mollinga P, and Karar M (2007) Policy and institutional reform: The art of the possible. In: Molden D (ed.) Water for Food, Water for Life: Comprehensive Assessment of Water Management in Agriculture. London: Earthscan. Moench M (2003) Groundwater and poverty: Exploring the connections. In: Llamas MR and Custodio E (eds.) Intensive Use of Groundwater: Challenges and Opportunities, pp. 441--456. Lisse, The Netherlands: Swets and Zeitlinger. Moench M (2007) When the wells run dry but livelihoods continue: Adaptive responses to groundwater depletion and strategies for mitigating the associated impacts. In: Giordano M and Villholth KG (eds.) The Agricultural Groundwater Revolution: Opportunities and Threats for Development, pp. 173--194. Wallingford: CABI. Molle F and Berkoff J (2006) Cities versus agriculture: Revisiting intersectoral water transfers, potential gains, and conflicts. Comprehensive Assessment of Water Management in Agriculture Research Report 10. Colombo: International Water Management Institute. Mukherji A (2006) Is intensive use of groundwater a solution to world’s water crisis? In: Rogers PP, Llamas MR, and Martı´nez-Cortina L (eds.) Water Crisis: Myth or Reality? Marcelino Botin Water Forum 2004, pp. 181–193. London, UK: Balkema/ Taylor and Francis Group. National Research Council (2004) Valuing Ecosystem Services. Washington, DC: National Academies Press. Neher PA (1990) Natural Resource Economics: Conservation and Exploitation. New York: Cambridge University Press. Ostrom E (1990) Governing the Commons: The Evolution of Self-Governing Irrigation Systems. Cambridge: Cambridge University Press. Peck JC (2007) Groundwater management in the high plains aquifer in the USA: Legal problems and innovations. In: Giordano M and Villholth KG (eds.) The Agricultural Groundwater Revolution: Opportunities and Threats for Development, pp. 296--319. Wallingford: CABI. Pongkijvorasin SP and Roumasset J (2007) Optimal conjunctive use of surface and groundwater with recharge and return flows: Dynamic and spatial pattern. Review of Agricultural Economics 29(3): 531--539.

Ragone SE and Llamas MR (2006) The alicante declaration: Steps along the pathway to a sustainable future. Ground Water Reader’s Forum 44(4): 500--503. Rijsberman F (2004) Sanitation and access to water. In: Lomborg B (ed.) Global Crises, Global Solutions, pp. 498--527. New York: Cambridge University Press. Rosegrant M, Cai X, and Cline S (2002) World Water and Food to 2025. Washington, DC: IFPRI. Rubio SJ and Fisher AC (1997) Adjusting to climate change: Implications of increased variability and asymmetric adjustment costs for investment in water reserves. Journal of Environmental Economics and Management 34: 207--227. Saleth RM (1996) Water Institutions in India: Economics, Law and Policy. New Delhi: Commonwealth. Selborne J (2001) The ethics of freshwater use: A survey. Report of the Commission on the Ethics of Science and Technology (COMEST), 62pp. Paris, France: UNESCO. Schlager E and Lopez-Gunn E (2006) Collective systems for water management: Is the tragedy of the commons a myth? In: Rogers P, Llamas MR, and Martı´nez-Cortina L (eds.) Water Crisis: Myth or Reality? pp. 43--60. London: Taylor and Francis. Shah T (2005) The new institutional economics of India’s water policy. In: International Workshop on African Water Laws: Plural Legislative Frameworks for Rural Water Management in Africa0 . Johannesburg, South Africa, 26–28 January 2005. Shah T, Burke J, Villholth K et al. (2007) Groundwater: A global assessment of scale and significance. In: Molden D (ed.) Water for Food, Water for Life: Comprehensive Assessment of Water Management in Agriculture, pp. 395–423. London: Earthscan. Stalgren P (2006) Corruption in the Water Sector: Causes, Consequences and Potential Reforms, Swedish Water House Policy Brief No. 4. SIWI. Transparency International (2008) Global Corruption Report 2008: Corruption in the Water Sector. Cambridge: Cambridge University Press. Tsur Y and Parker D (1997) Decentralization and Coordination of Water Resource Management. Boston: Kluwer. UNECE (1998) Convention on access to information, public participation in decisionmaking and access to justice in environmental matters. Aarhus, Denmark, 25 June 1998. http://www.unitar.org/egp/sites/default/files/aarhus_convention.pdf (accessed March 2010). Wagner M (2005) Wildlife and Water: Collective Action and Social Capital of Selected Landowner Associations in Texas, 150pp. PhD Thesis, Texas A&M University. Wegerich K (2006) Groundwater institutions and management problems in the developing world. In: Tellam JH, Rivett MO, and Israfilov RG (eds.) Urban Groundwater Management and Sustainability. Dordrecht: Springer. WHO (World Health Organization) (2001) Arsenic in Groundwater – Factsheet. http:// www.who.int/mediacentre/factsheets/fs210/en/print.html. (accessed March 2010).

Relevant Websites http://www.acaciawater.com Acacia Water. http://www.aueas.org Asociacio´n Espan˜ola de Usuarios de Aguas Subterra´neas. http://www.whymap.org BGR: UNESCO; World-Wide Hydrogeological Mapping and Assessment Program (WHYMAP). http://www.connectedwater.gov.au Connected Water: Managing the Linkages between Surface Water and Ground Water. http://www.ecolex.org ECOLEX: The Gateway to Environmental Law. http://www.empowers.info EMPOWERS Thematic Group: Advancing participation and dialogue in local water governance in the MENA Region. http://www.epa.gov EPA: United States Environmental Protection Agency; Groundwater and Drinking Water. http://www.ewater.eu eWater Portal. http://www.water.nstl.gov.cn GWRTAC: Ground-Water Remediation Technologies Analysis Center. http://www.waterandfood.org Groundwater Governance in Asia: Water is Divine. http://www.sg-guarani.org Guarani Aquifer System.

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http://www.indiawaterportal.org India WaterPortal. http://www.igrac.nl International Groundwater Resources Assessment Centre. http://www.inpim.org International Network on Participatory Irrigation Management. http://www.waterlaw.org International Water Law Project: Addressing the future of water law and policy in the 21st century. http://www.isarm.net isarm Internationally Shared Aquifer Resources Management; Transboundary Aquifers. http://aguas.igme.es No se encuentra la pa´gina. http://www.groundwatermanagement.org Participatory Groundwater Management. http://www.ploppy.net Ploppy (educational material for children (in Spanish) on groundwater).

http://www.projectwet.org ProjectWET; Worldwide Water Education. http://www.groundwater.org The Groundwater Foundation. http://www.worldbank.org The World Bank; GW-MATE. http://www.iah.org The WorldWide Groundwater Organisation: International Association of Hydrogeologists. http://www.water-ed.org Water Education Foundation. http://www.wfdvisual.com WFDVisual: Water Framework Directive Visualisation Package. http://www.groundwateruk.org UK Groundwater Forum: Raising Awareness of Groundwater.

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1.08 Managing Agricultural Water J Ramirez-Vallejo, Universidad de los Andes, Bogota, Columbia & 2011 Elsevier B.V. All rights reserved.

1.08.1 1.08.1.1 1.08.1.2 1.08.1.3 1.08.1.3.1 1.08.1.3.2 1.08.1.3.3 1.08.1.3.4 1.08.1.3.5 1.08.1.3.6 1.08.1.3.7 1.08.1.3.8 1.08.1.4 1.08.1.4.1 1.08.1.4.2 1.08.1.4.3 1.08.1.4.4 1.08.1.4.5 1.08.2 1.08.2.1 1.08.2.2 1.08.2.2.1 1.08.2.2.2 1.08.2.2.3 1.08.2.2.4 1.08.2.3 1.08.3 1.08.3.1 1.08.3.2 1.08.3.3 1.08.4 1.08.4.1 1.08.4.2 1.08.4.3 1.08.4.3.1 1.08.4.3.2 1.08.4.4 1.08.4.4.1 1.08.5 1.08.6 1.08.7 1.08.8 1.08.9 References

Introduction and Overview Trends in Water Management for Agriculture Water Scarcity: Is It a Demand or a Supply Problem? Challenges Facing Agricultural Water Management The policy and institutional challenge The economic and financial challenge The problem of declining investment The challenge of technology and water resources to supply growing demand Poor performance of public managed irrigation systems The neglect of environmental impacts of agricultural water management Neglect of water management for rainfed agriculture The poverty and rural incomes challenge Potential/Promise of Science and Technology Advances Hydro-climatic forecast prediction Drought-tolerant crops Remote-sensing technology on the estimate of crop ET Cost-effective irrigation systems Improving irrigation water management Water Productivity in Agriculture Economic Value of Water for Agriculture Example of Estimates of the Economic Value of Water: The Case of Mexican Agriculture Indirect method Residual method Water markets Math programming method Agricultural Trade Protection Water Management and Competitiveness Framework Farmers, Water, and the Process of Economic Development Water Resource Management and the Regional Economic Strategy Water Resource Management, Institutions, and Implementation Integrated Water Resource Management Participatory Irrigation Management Lessons Learned from Participatory Management The Philippines Mexico Institutions and Water Governance Water management principles Water Management and the Environment Water for Agriculture and Poverty Reduction Water Management of Rainfed Agriculture Policy Actions for the Future Summary

1.08.1 Introduction and Overview Managing water for agriculture is a topic that covers broad dimensions in the development of countries, such as management of water for rainfed agriculture, irrigated agriculture, water-use recycling, conservation of water and land, and watershed management. It is a collection of activities that

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lies between four areas of economic development: rural development, agriculture, water supply, and environmental management. Water management has achieved an impressive record during the past century but there is a larger challenge ahead. Irrigation is responsible for approximately 75% of water demand in developing countries and low-cost waters are already

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a scarce resource. Only 17% of all cropland is irrigated, but provides 30–40% of the world’s food production, and over 60% of the world’s irrigated area is in Asia, mostly devoted to the production of rice (IUCN and WBCSD, 2008). Water needs to be more productive as an input to generate a more dynamic development of rural areas, to supply the food for the increasing demand in the world, and to leave enough water for other key uses such as domestic and industrial uses. At the same time, water use in agriculture needs to be friendly with the environment and society in general. A new context for water management in agriculture has been generated by the changing national and global trends. The high rates of population growth, the dynamic economies of countries such as China, India, and East Europe that bring each year millions of people out of poverty with increasing levels of disposable income to spend on higher value food, and the stronger desire of the world to live in an environmentally safe planet are just some of the trends that delineate the context in which water will need to be managed for agriculture. Water connects all ecosystems across the landscape and the competition for its use is becoming increasingly intense over time. In many locations, water supply is limited compared to its demand for various uses, and given that agriculture is by far the largest consumer of water, the future effective supply will depend on the productivity of water in agriculture. This water productivity is a function of the management, innovation, and governance of the water resource systems. More efficient irrigation systems, more precision agriculture, and the upgrading of rainfed systems and wastewater management will generate a better social and economic water use when agricultural production is increased. The above scenario suggests the importance of dealing with water management in agriculture as a way to optimize its use for society. This chapter first presents the context and main trends that will influence water management in agriculture. It presents the main challenges facing agricultural water management, and its impact on poverty reduction. Subsequently, two related sections present the link between water management for agriculture and competitiveness as a framework to explore water productivity in agriculture, with emphasis on the value of water for irrigation. Topics such as water management trends, water management in rainfed agriculture, and the impact of water management on the environment are also covered. Finally, governance and institutional aspects of water management are delineated.

1.08.1.1 Trends in Water Management for Agriculture According to the United Nations (2008), it is estimated that by year 2050 the world population can reach 8.9 billion, up by 47% of today’s population, and will place a significant pressure to world agriculture to satisfy food demand. Global agriculture must double in the next 30 years to sustain the population, and given the constraint on land expansion, current agriculture will need to become more efficient to meet this demand. Given this reality, the obvious question is how the world is going to use the available water to produce enough food to feed the world population in the future without compromising the environment and with a minimum

external cost for society? Also, given that it is expected that the urban population will exceed the rural population by a significant margin, how can water be more productive to deal with a shrinking rural population that needs to feed increasing urban population as well as themselves? Although the share of world’s workers who are employed in agriculture has decreased from 42% in 1996 to 36% in 2006 (International Labour Organization, 2007), this sector will remain to be an important source of occupation for many people, especially those from developing countries. Water will remain with an important role of increasing the well-being of many rural workers and their families, directly through consumption of the fruits of their labor and by increasing their income. Consumers will define the function of water management in agriculture. As income increases so does the demand for food, not only in terms of quantity but also in the composition of the traditional food intake. It has been shown that wealthier consumers demand more animal protein, such as meat and milk products, something that requires more land to be used for agriculture and livestock production. As an example, in the recent evolution of diet in China, meat consumption has more than doubled in the last 20 years and it is projected to double again in 2030 (Centre for World Food Studies, 2008). Another observed trend is the concern of the consumer about food safety where traceability becomes important and water becomes an attribute of agricultural products when consumers take the final purchasing decision. In addition, people in developed countries will demand more fruits and vegetables, a trend that will induce the production of fewer calories per hectare, reducing field crops such as cereals and producing more vegetables and fruits that require more water. Food production to satisfy a person’s daily dietary takes about 3 m3 of water, a little more than 1 l calorie1 (IUCN and WBCSD, 2008). Another trend that will transform the composition of food intake is the change in consumer taste and expectations when people place more emphasis on doing their best with the limited time at their disposal. People are eating healthier convenient food such as fresh-cut snacks, a market that has grown from 8.8 billion in 2003 to 10.5 billion by 2004, according to the International Fresh-Cut Produce Association. These trends are becoming stronger over time. Water plays then an active role in the virtuous cycle of development. Water is needed to generate high-value products and income to rural people that will demand more agricultural products. In addition, more sophisticated products will create an incentive for innovation processes. The pressure to increase water productivity in agricultural ecosystems will induce water to be used in a more friendly manner with numerous species, such as microorganisms in the soil and pests and predators that are part of the agricultural biodiversity needed to have plant stability and sustainable crop production. Water is then a necessary input to implement management systems such as integrated crop management and conservation agriculture. Water pollution is increasing and the sources of water are being affected by the augmentation in water consumption mainly for agricultural production. Irrigation water has

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generated important increases in income in many parts of the world with a clear impact on poverty but with negative effects on the environment. The problems of soil erosion, salinization, intrusion of seawater, and pollution, in general, are limiting the effective water availability for agriculture and other uses. Most of these processes are real consequences of badly managed water resources in settings with institutions that fail or are slow to change processes. Land and water productivity had raised agronomic yields from 1.4 tons per hectare to 2.7 tons per hectare over the past four decades. However, according to the IWMI (2007), the number of malnourished people is above 850 million (IWMI, International Water Management Institute), and the average daily per capita food supply in some regions of the world, such as South Asia and sub-Saharan Africa, is rising too slowly to generate a significant impact on poverty alleviation. These levels are far below from the ones observed in industrialized countries. In addition, it is also estimated that a third of what it is produced in agricultural products is lost before it is consumed. Another trend that is set to play an important role in future water management in agriculture is the world balance of energy. It is estimated that alternative fuels will provide 5% of the United States’ energy by 2020, up from 1% today, and something similar will be experienced by other developed and developing countries. Bioenergy, currently made largely from sugar cane and from corn, will increase the demand for water and generate water-associated contamination. More certification standards and enabling regulatory and policy frameworks will most likely be set to obtain sustainable practice of this economic activity. Moreover, competition for land use to provide food and fiber will affect food prices and impact lowincome consumers. There is still some debate regarding the extent to which climate change will impact agricultural productivity at the global level (World Bank, 2008). Climate change is affecting precipitation patterns and temperatures, and it is estimated that the areas of the world that are poorer will be the most adversely affected in its impact on water supplies. Groundwater levels are declining in many areas of the world, such as North Africa, North China, India, and Mexico, because of overexploitation, and there is not enough water to meet all demands of the different uses. Some estimates show that if nothing is done, climate change by the 2080s would have an impact in agricultural production capacity by about 16% if carbon fertilization is omitted and by about 3% if it is included. Some parts of the world will be more affected than others, particularly India and a number of countries in Africa will be impacted. Finally, the reconfiguration of global value chains that redefine the role of location on generating advantages to companies is a trend that will modify the geography of agricultural output and production of goods and services that are used as inputs in agriculture.

1.08.1.2 Water Scarcity: Is It a Demand or a Supply Problem? When confronting the future, the relevant question is whether the relative water scarcity is a demand or a supply challenge. Some authors are optimists and some others are pessimists. The optimists, for example, Allan (2001), the World Bank

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(2003), IFPRI (1995, 1997), Dyson (1996), Rosegrant et al. (1995), Rosegrant and Cai (2002), and Brichieri-Colombi (2004, Who Speaks for the River; PhD Thesis, University of London; unpublished) think there is sufficient freshwater and soil water in the world to meet current and future water needs if we accept the demographic predictions. On the other hand, we find pessimists such as Postel (1999) and Postel and Richter (2003) who argue that there are not enough water supplies to meet the estimated demand, given the population and income projections. For them, the prospect looks darker when considering the water-scarce Middle East and the heavily populated South Asia and China. For some, the difficulty can be described as a technology and economic problem to make water available where it is needed for distinct uses. In other words, there is plenty of water in the world – the issue is how to induce the right costefficient technology to obtain the desired quality of water available to transport it to where it is needed. Theoretically, the existing pressure on freshwater use will generate the right incentive to innovate new ways of obtaining water, such as (1) the production of freshwater by desalination of brackish or saltwater (mostly for domestic purposes), and (2) the reuse of urban or industrial wastewaters (with or without treatment), which increases the overall efficiency of use of water (extracted from primary sources), mostly in agriculture, but increasingly in the industrial and domestic sectors. This category also includes agricultural drainage water. However, these technologies are still high-cost options for most developing countries, a fact that leads to reframe the water scarcity problem as a demand problem – whether economic resources are available at the country, region, company, and farm levels to pay for the cost of these technologies. The two major cost items are treatment and transportation to users. Most likely, in a scenario with increasing water demand accompanied by scarcity of local water supply, the induced innovation hypothesis will take place (Hayami and Ruttan, 1985), in which technical and institutional changes will be induced through the responses of farmers, agribusiness entrepreneurs, scientists, and public administrators to water endowments and to changes in the supply and demand of water and other inputs. Based on this, the state of relative endowments and the accumulation of the primary resources of land, labor, and water, would be critical elements in determining the pattern of technical change that will occur in water resources and agriculture in the future. Technical change embodied in new and more productive inputs may be induced primarily to save labor, to save land, or to save water. However, this works in the right direction if the agricultural inputs are priced according to their opportunity cost (Ramirez-Vallejo and Rogers, 2004). If subsidies exist, which is the case in the majority of developing and developed countries, water will remain undervalued by society and future changes in technology will not reflect the reality of water scarcity. A trade-related concept has been proposed to solve the water-supply–water-demand correspondence. The trade-in embodied water or virtual water concept comes from the idea that water should be treated as a production factor and is equivalent to the volume of water needed to produce a commodity or service. The virtual water argument, on the other hand, shows that the importation of agricultural products

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that require significant amounts of water represents the importation of water into a water-scarce country. According to Allan (2004), ‘‘trade in commodities has already achieved a pivotal position in enabling the comparative advantages and disadvantages of regional water endowments to be balanced.’’ The concept of trade-in embodied water can be used to study the linkages that exist between trade reform and water resources development. The trade-in embodied factor or the factor content of trade approach was first employed by Leontief (1953) in his well-known test of the Heckscher– Ohlin (H–O) theorem and later formalized theoretically by Travis (1964), Melvin (1968), Deardorff (1982), and many others. However, there are only few studies in the literature that discuss trade-in embodied environmental factors. This approach has been extended to analyze the linkage between trade, food security, and water resources by using the virtual water concept and argument. The concept originates from the idea that water should be treated as a production factor and virtual water is the volume of water needed to produce a commodity or a service. In other words, countries sometimes do not have a surplus amount of water to produce water-intensive agricultural commodities such as rice, cotton, and tropical and subtropical fruits such as bananas and oranges, and use trade to compensate for this shortage. Food trade then becomes an instrument to augment water supplies on the scale needed to meet the domestic food demand. This concept has become popular since Allan (1996) used it first. An example of this concept is that it takes approximately 1500 m3 of water to grow 1 ton of grain. Equivalent figures can be estimated for virtually any traded agricultural product; however, these figures can only be used as an approximate number given that the real demand and supply of water to an individual crop depend on a large number of variables (i.e., the geographic location of the region where the crop is grown, the irrigation system used, the management of the irrigation system, soil type, climatic and socioeconomic conditions, etc.). The concept of virtual water then becomes a useful indicator of future water resources use. The virtual water argument can be also seen as an application of the well-known H–O theorem. The assumptions of this model are that factors can move without cost among industries within a country, but are completely immobile internationally, which might be the case of water as a production factor. As such, production functions for all countries (products) exhibit constant return to scale and each country has the same productive technology for each good, and consumers have the same utility functions. Under these conditions, the theorem states that a country exports products that use intensively a relatively abundant input. However, the level of protection to agricultural products via tariffs and duties and nontariff instruments by all countries, developed and developing, has distorted the virtual water movement worldwide. Ramirez-Vallejo and Rogers (2004) showed that, for example, using International food Policy Research Institute (IFPRI)’s IMPACT model results, a scenario of full liberalization of agriculture compared to a baseline scenario would have a significant net effect of virtual water flows from the relocation of meat and cereals trade. When the net effect of the meat and cereals markets are added together,

the two major contributors to the increase in virtual water trade would be the United States, which would increase its annual virtual water exports in about 86 km3, and Latin America would have a similar increase of 89 km3. These become the two water surplus regions in the world. The major changes in virtual water imports would occur in Asia in general (South Asia, Southeast Asia, and East Asia) with an increase of 112 km3, sub-Saharan Africa with an increase of almost 40 km3, and the former Soviet Union with an increase in water imports of 22 km3, mostly because of an increase in meat imports. West Asia and North Africa together, on the other hand, would decrease the level of virtual water imports to about 7 km3, but would remain as an important net importer of virtual water of about 176 km3. Ramirez-Vallejo and Rogers (2009) showed that the concept of virtual water is useful to educate public officials and society in general that water in some parts of the world is a scarce resource and that agriculture uses the great majority of water resources available on earth. The argument also has an implicit lesson underscoring the importance of running efficiently irrigation districts so that water could be allocated to other uses, including ones benefiting the environment.

1.08.1.3 Challenges Facing Agricultural Water Management Under the described trend scenarios, agricultural water management faces many challenges, especially as a result of its strong links with many global and economic issues. Agricultural management is not a unique goal by itself but it is part of a process to manage one of the most important inputs to income generation. According to the World Bank (2006), some of the challenges facing agricultural water management include the policy and institutional challenge; the economic and financial challenge; the problem of declining investment; the challenge of technology and water resources to supply growing demand; the poverty and rural incomes challenge; and the environmental dimension and sustainability imperative.

1.08.1.3.1 The policy and institutional challenge The most difficult challenge has been how to conciliate agricultural and macroeconomic and social policy. Countries want low-cost food products for the population and at the same time to improve farmer’s income levels. Low-cost products for local consumers imply low trade barriers and higher competition from international competitors, and lower support levels to agriculture for a food security policy. Generating incentives for agricultural development through water management for agriculture has its direct trade-offs with specific macroeconomic goals. The incentives for agricultural water management need to be integrated with agricultural policy and new institutional capacity needs to be built for a better water allocation and priority setting at the basin, regional, and national levels. New management skills and a better understanding of the political economy of reforms need to be developed.

1.08.1.3.2 The economic and financial challenge Use of water in agriculture generates the lowest value-add compared to other uses such as municipal, domestic, and industrial uses. When water is assigned to the higher value use, agriculture is usually ranked lower and other benefits different

Managing Agricultural Water

from direct economic and financial need to be added to justify allocation. Water scarcity increases and competition creates the incentive to improve the returns on water. The economic challenge is how to build an incentive system to encourage efficient water use and profitable agriculture of high value. A broader economic challenge deals with the generation of an incentive framework for all types of investment and to promote environmentally responsible use. On the financial side, the challenge is to achieve cost recovery in traditional agricultural systems that are characterized by high subsidies, selfsufficiency goals, and low-productivity agriculture and to create the investment environment under the existing distorted incentive frameworks.

1.08.1.3.3 The problem of declining investment The global irrigated area doubled from 1960 to 2000, developing faster at the beginning of the period and slowing down in later years (Cleaver and Gonzalez, 2003; Winpenny, 2003). The world experienced a decrease in the construction of dams for the development of surface irrigation, which has been compensated for by the growth of groundwater irrigation. According to the World Bank (2005), governments are investing less in agriculture worldwide; public investment in agriculture has dropped, and investment in irrigation, drainage, and other agricultural water management projects has also been declining worldwide. The situation for groundwater irrigation is delicate. Most of the past investment has been predominantly private and done in an unregulated environment ending up in overexploiting groundwater resources. Another area that demands significant resources for investment is drainage. Unfortunately, it is difficult to provide the right investment incentives for drainage infrastructure because usually the benefits are underestimated and the cost recovery is a difficult task.

1.08.1.3.4 The challenge of technology and water resources to supply growing demand Irrigated agriculture supplies close to 40% of the world’s food and occupies only 17% of the cultivated land. However, for the future, the Food and Agriculture Organization (FAO) estimates that by 2030, food production needs to grow at 1.4% a year, and about half of this growth would have to be generated from irrigated agriculture. Therefore, the main challenge has to do with the capacity of the agricultural sectors to meet this additional demand. Technological change has slowed down from the significant advances of the Green Revolution of the 1960s, and the water base is overexploited in many ecosystems, particularly in developing countries where institutions and regulations failed to control water exploitation. However, technology is available, but often is not disseminated and adopted. Cost incentives or profitable market opportunities generate the incentive for farmers to invest in technology. A study by the World Bank found that there is ‘‘more technology available than we know what to do with’’ (World Bank, 2005). Many innovations are available to improve productivity or conserve water but have not been adopted as it was expected, such as drip technology, for example, that has been adopted on less than 1% of irrigated lands worldwide with affordable costs.

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1.08.1.3.5 Poor performance of public managed irrigation systems Publicly managed irrigation systems account for half the irrigated area in developing countries and their performance has been generally below their technical and economic potential. Water service is poor, unreliable, and untimely. Among the causes are the decline of agricultural prices and subsidies to basic crops that have hampered diversification to high-value crops. However, perhaps the major cause comes from poor institutions with weak organization structures, bureaucratic with inefficient top-down approaches to service. Deficient management has led to lower collection rates of operation and insufficient maintenance fees, which have led to higher needs for rehabilitation intervention, which creates a never-ending pervasive cycle. Scarce farmers’ participation in the management processes has been recognized as one of the main reasons for the lack of accountability and efficiency in the public institutions.

1.08.1.3.6 The neglect of environmental impacts of agricultural water management As a result of agriculture being by far the largest user of land and water resources, there is a long list of environmental costs and risks of using water for agriculture, including land degradation, salinization, and erosion; loss of environmental water flows; pollution; destruction of natural habitats and livelihoods through drainage of wetlands and through land expansion and deforestation; and waterborne disease. Drainage was neglected in the rapid expansion of irrigation and irrigated land has become waterlogged and salinized due to the rise of the water tables and accumulation of salts, becoming a constraint to productivity. Land degradation caused by agricultural water management practices and by lack of drainage is affecting some of the world’s most fertile basins and dams, and irrigation infrastructure modifies flows, affecting their seasonality and frequency of floods. In addition, irrigated agriculture is a source of pollution, as a result of the technology package promoted with the green revolution that included the important use of chemical inputs. However, the main challenge of the management of water in agriculture is to simultaneously conserve biodiversity and diminish any external impact of water use to increase production, to secure enough production to meet the increasing demand for food, and, finally, to improve the prosperity of the rural people all around the world. Management of water in agriculture needs not only further increase in the productivity of existing farmland, but also to do it simultaneously with a support of biodiversity and ecosystem services and efficiently manage the natural resources. It needs to foster healthy populations and help them realize their development potentials and to increase prosperity in general, by generating income and improving the livelihoods of the rural communities and increase the value of the agricultural products.

1.08.1.3.7 Neglect of water management for rainfed agriculture Rainfed areas are home to most of the poor habitants of the rural areas and account for 60% of the current agricultural output. These areas are characterized by having less use of

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available technology. The technological options for improved water management in rainfed agriculture are few, mainly because the incentives have not generated the relevant research results around the world to offer higher yielding varieties for water stress and environmentally sensitive conditions. As a result, improvements in water management have been limited in these areas.

1.08.1.3.8 The poverty and rural incomes challenge Agricultural growth is central to poverty reduction. Seventy percent of the world’s poor live in rural areas, and most of them are dependent on agriculture. Typically, the rural poor live on marginal lands or on drylands, with little or no access to controlled water sources. Their technological options for improved water management are limited, and they face high risks from rainfall variations. The poor are exceptionally vulnerable to drought, floods, effluent discharge, aquifer depletion, waterlogging, salinization, and water quality deterioration. Thus, the key agricultural water challenges for the poor are food security, risk mitigation, and income growth. One of the Millennium Development Goals (MDGs) – eradicate extreme poverty and hunger – can be achieved only if agriculture grows and can provide access to food for the poorest and most vulnerable. Improved management of available water thus has a critical role to play in poverty reduction and food security. Other MDGs such as gender equality, child nutrition, and market access also depend directly or indirectly on pro-poor agricultural growth and related management of scarce water.

1.08.1.4 Potential/Promise of Science and Technology Advances At the same time, there are challenges, some of which might be more optimistic concerning agricultural water management because of the opportunities provided by advances in science and technology. Some of them are the hydroclimatic forecast and prediction, which could lead to water saving and profit gain, remote-sensing technology on the estimate of crop evapotranspiration (ET), drought-tolerant or salt-tolerant crops, improvement of cost-effective irrigation systems, and best management practices, mainly for water quality. Some of these are explained below.

1.08.1.4.2 Drought-tolerant crops Drought conditions are a well-known restriction to crop production. Research in public and private institutions is producing promising results in terms of developing varieties that demand less water while maintaining a higher level of agronomic yield. Local plant biologists are working hard to develop new crop strains that will produce abundant food even when water is scarce. Canola, in Canada, for example, is highly sensitive to water stress, which is one of the main factors that limits crop yield (Wanna et al., 2009). Development of drought-tolerant canola was then considered an important and urgent mandate for the canola industry. A breakthrough in this area should increase yield and permit expansion of canola growth regions. However, to be an effective solution, the new strain must produce normal yields under nondrought conditions and produce greater yields than conventional varieties under stress conditions. Research efforts in order to create drought-tolerant crop plants have been intensive, but the results remain incremental. Monsanto and other agricultural companies have drought-tolerant soybeans, cotton, and other crops in the pipeline. Universities are also engaged in intensive research aimed at growing more food with less water. ‘‘A more rapid growth in rainfed yield and production could compensate for reduced investments in irrigation or reduced groundwater pumping to eliminate groundwater overdraft, but that achieving the required improvements in rainfed production would be a significant challenge’’ (Rosegrant et al., 2001).

1.08.1.4.3 Remote-sensing technology on the estimate of crop ET Spatiotemporal information on actual ET helps users to better understand evaporative depletion and to establish links between land use, water allocation, and water use. Satellite-based measurements, used in association with energy-balance models, can provide the spatial distribution of ET for these linkages and optimize the use of water for irrigation (Bastiaanssen et al., 2005). Remote-sensing and hydrological models are applied to irrigation projects to estimate the water balance to support water use and productivity analyses. A case study in the Yakima River basin (Washington State) demonstrates how ET from remote sensing can be used for evaluating water-conservation projects.

1.08.1.4.4 Cost-effective irrigation systems 1.08.1.4.1 Hydro-climatic forecast prediction Expert system techniques have been used in irrigated agriculture to optimize the use of water in districts and to maximize profits by increasing agronomic yields. In the Havana Lowlands region, Illinois, USA, for example, researchers incorporated different types of weather forecasts into irrigation scheduling for corn production. The results showed that if farmers just use the real-time soil moisture information and the empirical rules set, profits might increase by 16%; and over the five testing years, it was found that the proper use of the 7-day forecast could save irrigation water and increase crop yield in dry years. The study found that farmers at the study site sometimes applied more water than necessary (Bastiaanssen et al., 2005).

Water is also saved through innovation in irrigation systems. Cost-effective irrigation systems allow for a better allocation of water resources. One way to achieve this is through the reengineering of the irrigation systems, which consists of designing the most cost-effective answer to the redefined water service within the scheme. This is a solution that demands consideration of the spatial distribution of the effective demand for the water service and the spatial distribution of the physical infrastructure characteristics. Some of the people in rural areas have developed their own water system with cost-reducing adaptations. One of these is the development of a solar system, which requires low operation cost compared to an electrical system to drive the motors or engines that finally converts the electrical energy to

Managing Agricultural Water

mechanical energy, so as to pump the water to the required destination.

1.08.1.4.5 Improving irrigation water management Best management practices are also a way to save water in irrigation systems. Improved performance in irrigation water management can usually be achieved through rehabilitation, process improvement, which consists of intervening in the process without changing the rules of the water management, and modernization. The introduction of modern techniques is a process improvement, and modernization, which is a more complex intervention implying fundamental changes in the rules governing water resource management. It may include interventions in the physical infrastructure as well as in its management.

1.08.2 Water Productivity in Agriculture Water scarcity can be seen under a multidimensional framework of physical, economic, managerial, institutional, and political water scarcity (Molle and Mollinga, 2003): physical water scarcity in which water availability is limited by natural availability; economic water scarcity when human and financial resources constrain availability of water; managerial water scarcity where availability is constrained by management limitations; institutional water scarcity where water availability is constrained by institutional shortcomings; and political water scarcity where political forces bar people from accessing available water resources. These types of scarcity can occur concomitantly, increasing both the severity and the impacts of water scarcity. Molden et al. (2003) estimated that, by 2020, approximately 75% of the world’s population will live in areas experiencing physical or economic water scarcity, precisely where most of the poor and food-insecure people live. Water productivity is a concept that has meaning and use when water is a scarce resource. Higher water productivity becomes part of the answer to the challenges described above. However, productivity is a concept used frequently by economists, engineers, and even biologists, with some degree of confusion. The concept of water productivity relates to the desire to have a higher level of output, economic value-add, or just value for the aggregate production per unit of water used. In other words, it is the way of achieving more using scarce water resources. In order to better understand the concept of water productivity, it is necessary to review the definition of efficiency. Irrigation efficiency is defined as the ratio of water consumed to water supplied. Irrigation systems in the developing world typically function at a level of irrigation efficiency from 30% to 40% (Seckler et al., 2003), showing the big difference that exists in terms of quantity between the point of water diversion in the irrigation system and the water available in the root zone of the plant. Agriculture depletes the water resources through evapotranspiration, and this signifies less water for other uses in the system. Climate plays an important role in the total amount of water that is evaporated and transpired by plants.

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Although there is not a single definition, water productivity is then the ratio of crop output to water either diverted or consumed, being expressed in either physical or monetary terms, or some combination of the two (Barker and Molden, 2003). Another related concept used by economists is the economic efficiency, which takes into account value of output, opportunity costs of inputs and externalities, and it is achieved when water is allocated and used so the net value or net returns are maximized. It is a criterion that describes the conditions that must be satisfied to achieve the largest possible net resource (Wichelns, 1999). It is a concern for economists that, in general, the true value of water is not reflected by the prices or charges for irrigation water, in which the allocation does not follow socially optimal scenarios. Management of water in agriculture should be targeted to allocate water to the highest level of water productivity as a social optimum. In theory, the allocation of water among competing uses involves an economic optimization exercise. If we were to consider only private returns to the agricultural production activity, some alternatives to improve water productivity would be (Wallace and Batchelor, 1997):

• • • •

agronomic improvements (e.g., improved crop husbandry, cropping strategies, and crop varieties); technical improvements (e.g., improved and lower-cost technologies for extracting groundwater); managerial improvements (e.g., improvements in farmlevel resource management or system operation and maintenance (O&M)); and institutional improvements (e.g., introduction of water pricing and improvement in water rights).

The first two categories are innovations or new technologies that lower cost or increase the value of output per cubic meter of water. The third category refers to an increase in technical efficiency, and the fourth relates to allocative efficiencies encouraged by the creation of market incentives (Barker and Molden, 2003). The economic value of water – economic water productivity – can be increased by different actions: (1) increasing the agricultural yield per unit of water used; (2) switching from low value to high-value crops; (3) relocating water from low to higher value water uses; (4) lowering the cost of inputs for the same water used; (5) increasing health benefits and the value of ecological services of agriculture; (6) deceasing the social, health, and environmental costs; (7) obtaining multiple benefits per unit of water; and (8) achieving more livelihood support per unit of water, such as, more jobs, nutrition, and income. However, economic theory shows that if a new practice offers net benefits and does not have any negative effects on third parties off the farm (termed externalities by economists), then the adoption of this practice is advantageous for the society as a whole, not just for the farmer. Unfortunately, there are many externalities in the agricultural sectors of developing and developed countries that make it more complicated to identify the optimal practices that should be adopted by private agents (farmers).

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Policy-induced incentives are necessary to achieve an optimal water allocation. In order to assess the optimality of the allocation of water as a scarce resource, it is necessary to value all agricultural inputs and outputs at social prices that generally diverge from market prices. The difference between these two prices is a result of distortions, subsidies, and transfers among different actors of the economy.

1.08.2.1 Economic Value of Water for Agriculture The value of water is the maximum a user is willing to pay for the resource. In other words, water has value when users are willing to pay a price instead of not having it. Water has economic value when its supply is constrained compared to the amount that is demanded, and when water as an input becomes a constraint to the economic and social development process. This concept has become important particularly after water was declared as an economic good in Dublin, 1992. Water needs to be evaluated for the benefit of humanity, and to ignore its value would lead to an overexploitation of the resource and an incorrect allocation among alternative uses. Rational decisions that support water development, its allocation, and final use, demand a measurement effort to value water for alternative uses. When defining investment priorities in a country or a region, the level and the geographical variability of the economic value of water become a key element in the decision process. It is used as a selection criterion when compared to what users are actually paying for the resource (cost recovery and operation and maintenance fees). It also becomes a good indicator of the future demand of the resource for irrigation. There exists a set of methodologies to estimate the value of water for agriculture: direct methods such as the water markets and the contingent evaluation approach. There are also indirect methods such as the estimation of demand equations and its integration, and a series of methodologies that are derived from considering water as an intermediate good. Among these are the residual value methodology, the alternative cost, and the Hedonic price methodologies. Finally, when primary information does not exist to estimate the value of water, a methodology called meta-analysis is employed to extrapolate economic values of water from other watersheds or regions to the region of interest.

1.08.2.2 Example of Estimates of the Economic Value of Water: The Case of Mexican Agriculture Mexico is an interesting country case to observe the application of alternative methodologies to estimate the value of water for irrigation. Ramirez-Vallejo and Rogers (2004) used an indirect methodology proposed by Moore (1999) to estimate the value of water for irrigation, and also applied a residual method in various irrigation districts in Mexico. Others estimated direct values from water markets and the use of optimization exercises of the producer’s behavior at the farm level. Kloezen and Garce´s-Restrepo (1998) documented rent transactions of water concessions, and Florencio-Cruz et al. (2002) applied math programming to estimate values of irrigation water. As seen below, the estimates of the economic value of water for irrigation vary significantly depending on

the methodology employed, a situation that complicates the application of water value as a criterion for allocating water.

1.08.2.2.1 Indirect method The indirect methodology employed by Ramirez-Vallejo and Rogers (2004) was justified by the fact that the water prices are divorced from the production costs in the long run. This is something that has been criticized in the past with the argument of efficiency (Ciriacy-Wantrup, 1954; Bain et al., 1996). For the case of Mexico, a revenue function was defined that related the multiproduct to fix inputs, estimated as

Rðp; x; zÞ ¼ maxfpyA Yðx; zÞ; p4 0g

ð1Þ

where p is the vector of product prices and y is the vector of product production for each product, x is the amount of water, z is the amount of the mix input, and Y(x,z) is the set of production probabilities. Chambers and Just (1989) developed the properties of the revenue function, which was estimated econometrically, and the shadow prices were derived using a quadratic normalized form suggested by Lau (1978) to estimate the willingness to pay. Applying this methodology for 14 irrigation districts using data from 1993 to 2001 resulted in the following shadow prices for irrigation water (see Table 1).

1.08.2.2.2 Residual method Using an alternative residual method, Ramirez-Vallejo and Rogers (2004) computed the value of the production factors different from water and then subtracted this value from total sale income from cultivated products in the irrigation districts. This difference was assigned to water as its economic value. Young (1996) clarified that this imputed value is valid if two Table 1

Shadow prices for irrigation water ($ m3)

Irrigation district

Average shadow price 1993–2001

Average shadow price 1997–2001

001 Pabellon, Ags. 005 Delicias, Chih. 010 Culiacan Y Humaya, SIN. 011 Alto Rio Lerma, GTO. 014 Rio Colorado, B.C. 017 Region Lagunera, DGO. Y C. 023 San Juan Del Rio, QRO. 024 Cienega De Chapala, MICH. 038 Rio Mayo, SON. 041 Rio Yaqui, SON. 075 Rio Fuerte, SIN. 076 Valle Del Carrizo, SIN. 92A R.Panuco, Tamps. ‘‘ANIMAS’’ 92B R.Panuco Pujal Coy, S.L.P.

1.160 0.844 1.455

1.568 0.929 1.557

0.630 1.649 1.572

0.888 1.132 2.050

1.856 1.691

2.000 1.578

0.810 0.695 0.190 1.341 1.679

0.677 0.907 0.987 1.737 2.637

1.073

1.329

From Ramirez-Vallejo J and Rogers P (2004). Virtual water flows and trade liberalization. Journal of Water Science and Technology 49(7): 25–32.

Managing Agricultural Water

conditions are satisfied: first, the inputs and products are in competitive markets and are not regulated, that is, the price is equal to the value marginal product. Second, the production function should behave in such a way that an increment in each of the inputs generates an equal relative increment in the product. The residual value of water for irrigation was estimated by

Residual value of water ¼ PQ 

X

Wi Ni

where P is a vector of prices of products in the district and Q is a vector of the amount produced. Wi is the quantity of input i used and Ni its price. Besides the traditional inputs (land, capital, and labor), other inputs were considered such as the administration and the expected utility of the economic activity. The shadow price was then the residual value divided by the amount of water used. The results of applying this approach to some irrigation systems in Mexico are shown in the Table 2.

1.08.2.2.3 Water markets When water markets exist, prices applied in the resource transaction become good proxy of the economic value of water. Furthermore, if the market allows observing the resource–demand curve, the value would be determined by the area under the curve. There exist some water markets in irrigation districts in Mexico where some transactions involve the water concessions and others the right to use water for specific time periods. Table 2

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Kloezen and Garce´s-Restrepo (1998) documented the rent transactions of the water concessions for the irrigation district of Alto Rio Lerma (Table 3). The observed market prices were below the economic values estimated using alternative methodologies, a situation that could be explained by the distortions of rental markets of concessions in the irrigation districts as a result of direct intervention of user associations; social and political motivations were present in the exchange process of water rights, which did not allow water markets to function properly. Other registered transaction values of concessions for the irrigation district of Lagunera varied from $0.05 to $0.1 m3. Finally, the government, in its Aquifer Program, rationalized the use of water by purchasing some of the users’ concessions. In 2004, the government, using market mechanisms, purchased back concessions at an average price of $2.5 m3, equivalent to an annual value of $0.25 m3 after applying the perpetuity formula with a discount rate of 10%.

1.08.2.2.4 Math programming method Some authors have implemented the math programming method to estimate the economic value of water, optimizing the net income subject to various constraints, among them, the budget constraint, and the water availability. The execution of these exercises for different levels of water available allows estimating the water demand equation, which, when integrated, gives the economic value of water, or willingness to pay for the resource. Florencio-Cruz et al. (2002) executed this exercise to the Alto Rio Lerma irrigation district and a significant dispersion of the estimate was found (Table 4), depending on the season considered.

Shadow price using the residual method

1.08.2.3 Agricultural Trade Protection Typology

Irrigation district

(Shadow price Pw)

Rehabilitation New New New New New

Yaqui Yaqui Angostura Carrizo Aguascalientes Queretaro

0.159 0.242 0.879 0.394 0.151 0.212

From Ramirez-Vallejo J and Rogers P (2004). Virtual water flows and trade liberalization. Journal of Water Science and Technology 49(7): 25–32.

Table 3

Agriculture is one of the most important sectors in many developing countries in terms of social and political stability, and thus it is heavily circumscribed by global trade protection, much of it originating from Organization for Economic Cooperation and Development (OECD) countries. Agricultural policies in OECD countries cost consumers and taxpayers over $280 billion every year (Anderson et al., 2006). The value of total agricultural support in OECD countries was more than 5 times higher than total spending on overseas development

Rental price in the irrigation district of Alto Rio Lerma

Season

Seller

Buyer

Volume (m3  1000)

Bought as a % of total used

Price ($m 3)

Summer 1995

Acambaro Acambaro Acambaro Acambaro Acambaro Acambaro Valle Valle Jaral

Cortazar Salvatierra Huanimaro Salvatierra Abasolo Huanimaro Salamanca Abasolo Salamanca

10 000 10 000 2000 8000 3000 2000 3500 1000 450

25 21 23 36 19 86 18 6 2

0.004 0.009 0.009 0.020 0.034 0.034 0.035 0.034 0.035

Summer 1996 Summer 1997

From Kloezen W and Garce´s-Restrepo C (1998). Assessing Irrigation Performance with Comparative Indicators: The Case of the Alto Rio Lerma Irrigation District Mexico. Colombo, Sri Lanka: IWMI.

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Table 4 Value of irrigated water for Alto Rio Lerma District using math programming method Groundwater February April Surface water January May

0.7–1.8 $ m3 1.6–2.4 $ m3

•

•

3

0.5–1.2 $ m 1.3–1.9 $ m3

From Florencio-Cruz V, Valdivia-Alcala´ R, and Scott CA (2002) Productividad del agua en el Distrito de Riego 011, Alto Rio Lerma. Agrociencia 36: 483–493.

• assistance and twice the value of agricultural exports from developing countries (OECD, 2001). Reductions in distortions to international trade of agricultural products will most likely increase real income and stimulate change in the composition and location of production and consumption of water, as a primary production factor. The impact of agricultural trade on water resource development will differ between countries depending on how protected and distorted their agricultural sectors are as well as their water resources endowment, its use, and their policy and institutional framework. That is, where trade liberalization leads to a decrease in agricultural production, there is likely to be a reduction of water use and its associated environmental impact, but where it increases production, water will be allocated to agriculture from other actual or potential uses (Ramirez-Vallejo and Rogers 2009). Therefore, a change in trade barriers and agricultural subsidies in general could impact differently water allocation, and, therefore, water policy adoption. In theory, to achieve a rational decision to allocate water within alternative uses, it is important to measure the value of water in these uses. When social and market prices are equivalent and no distortions are present, the market as a clearing mechanism yields the optimum allocation of water for a society as a whole. When this is not the case, the externalities should be taken into consideration for the analysis. However, in practice, there is often little time, money, knowledge, or will to conduct serious economic analyses of benefits and costs to consider in the water allocation disjunctives. To summarize

•

•

• •

The concept of water productivity relates to the desire have a higher level of output, economic value added, or just value for the aggregate production, per unit of water used. In other words, it is the way of achieving more using less scarce water resources. Water productivity is then the ratio of crop output to water either diverted or consumed, being expressed in either physical or monetary terms, or some combination of the two. Management of water in agriculture should be targeted to allocate water to the highest level of water productivity as a social optimum. Some alternatives to improve water productivity are: (1) agronomic improvements; (2) technical improvements; (3) managerial improvements; and (4) institutional improvements.

The value of water is the maximum a user is willing to pay for the resource. Water needs to be evaluated for the benefit of humanity, and to ignore its value would lead to an overexploitation of the resource and an incorrect allocation among alternative uses. There exists a set of methodologies to estimate the value of water for agriculture: direct methods such as the water markets, and the contingent evaluation approach; indirect methods such as the estimation of demand equations and its integration; and a series of methodologies that are derived from considering water as an intermediate good. The impact of agricultural trade on water resource development will differ between countries depending on how protected and distorted their agricultural sectors are as well as their water resources endowment, its use, and their policy and institutional framework.

1.08.3 Water Management and Competitiveness 1.08.3.1 Framework Water management is intrinsically linked to the competitiveness upgrading process. Under the current global scenario, the goal of many countries and regions is to increase the level of competitiveness to achieve a higher level of prosperity. Regions and countries are competing to offer a more competitive environment to allow agricultural firms and farmers to be more productive. The opportunity that farmers and firms have of obtaining the right amount of water at the right time to optimize the value-generation process becomes an important characteristic of the business environment. However, the link between water management and competitiveness is not well understood and most countries assign higher priority to other types of policies. Developing countries focus most on reforms to adjust their macroeconomic, political, legal, and social conditions that are necessary conditions, but far from being sufficient to increase prosperity. Wealth is created by the productivity of all production factors that a nation or region can utilize, and heavily depends on the microeconomic conditions, understood as the elements that are located outside the production unit (i.e., farms) and the conditions that take place within the boundaries of companies and farms. Unless these microeconomic capabilities improve, the growth process of prosperity will be truncated (Porter, 1998). Competitiveness is the value that firms can generate per unit of input used, that is, capital, land, labor, and so on. This real value is created by the private sector, by the farmers themselves, when transforming these inputs into products and services that consumers are willing to pay for. A farm is more productive when it generates the same amount of output but the production is sold in the market for a higher price. In addition, a farm is more productive if it generates more physical output per unit of input. True competitiveness is then synonymous with productivity, and the higher the level of productivity of the agricultural units, the higher the wages they can support, and the more important the influence on prosperity. In the agricultural context, wealth is created at the microeconomic level when farmers and firms produce valuable

Managing Agricultural Water

goods and services using water more efficiently. The wealth creation process is a result of their production activity and not a direct result of government or other societal institutions’ intervention. Water improves farm productivity in various ways. First, it creates a direct impact on the possibility of technology adoption by farmers and agricultural firms. Without water, for example, the choice of agricultural varieties would be limited. Water also allows the use of other agricultural inputs such as fertilizers, and causes a clear impact on production stability that helps to generate a higher output value. Moreover, paradoxically, water scarcity also generates the right incentive for innovation, a phenomena that explains new technology advances available for rainfed agriculture. Under a water management for agriculture framework, the foundations of productivity rest on two interrelated areas: (1) the sophistication and capabilities with which farms and agricultural companies compete, which are highly dependent on the availability of water resources and (2) the quality of the business environment in which they operate. More productive farm and firm operating practices require highly skilled people, technology and availability of water resources, better information, more efficient government processes, improved infrastructure, better suppliers, more advanced research institutions, and more intense competition, among other things. The competitiveness of a farmer and an agricultural company, then, depends on both, their internal capabilities and the characteristics of their location, which are influenced directly by water management. Water is not only a regular input of the agricultural production and transformation process, but also a facilitator of the strategic upgradation of farmers and agricultural firms. Agricultural firms and farmers must upgrade their modes of competing and capabilities in order to generate economic

development. They need to shift from competing on costs and inherited endowments to create competitive advantages from efficient and distinctive products and internal processes. Competition on cost structure or price is usually the traditional way of competition within agricultural sectors in most countries, and a characteristic of earlier stages of development. Moving to more sophisticated ways of competing depends on parallel changes in the business environment. The business environment can be understood in terms of four interrelated areas: the quality of factor (input) conditions, the context for firm strategy and rivalry, the quality of local demand conditions, and the presence of the related and supporting industries. Because of their graphical representation (see Figure 1), the four areas have collectively become referred to as the diamond (Porter, 1998). Water management impacts directly the quality of the business environment for agricultural development. First, water improves the access to high-quality agricultural inputs, such as infrastructure, fertilizers, human resources, and capital. Second, water user associations usually generate rules and incentives that affect investment and productivity, and foster or inhibit a vigorous local competition. Third, advanced water management practices induce the sophistication of local customers and needs for example, strict quality, safety, and environmental standards. Fourth, improving the availability and reliability of water supply facilitates the existence of suppliers and supporting industries of agricultural development. Improving competitiveness is a special challenge because no single strategy or action of an individual institution can improve significantly the productivity of a region. Many things matter for competitiveness. The quality of schools and roads, the financial markets penetration, the consumer sophistication, the supplier networks, the rules for competition, the quality of the private and public demand, and many more

Context for firm strategy and rivalry

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• Water user associations could generate rules and incentives that encourage investment and productivity, and foster vigorous local competition

Factor (input) Conditions

Demand conditions

• Water improves the access to highquality agricultural inputs

• Advanced water management

Related and supporting industries • Improving the availability and reliability of water supply facilitates the existence of suppliers and supporting industries of agricultural development

Figure 1 Water management and quality of the business environment for agricultural development.

practices induce the sophistication of local customers and needs –e.g., strict quality, safety, and environmental standards

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attributes act as determinants of competitiveness of the regions where agricultural products and services are produced. Water resource management is one of them and needs to be articulated with reforms in other areas to upgrade competitiveness.

1.08.3.2 Farmers, Water, and the Process of Economic Development Improving competitiveness is a collaborative process in which private sector, academia, government, institutions, and labor unions contribute in designing and executing a competitiveness strategy. Farmers, entrepreneurs, and managers of agricultural companies play a crucial role in improving competitiveness and defining regional policy. Governments play an important role in upgrading competitiveness because it affects all its determinants. Government agencies affect regulatory standards, invest in public goods, set the rules and incentives, coordinate policies, and have an effect on the quality of its purchases of goods and services. Universities and schools impact the role of water in improving competitiveness through knowledge and human resource skill formation to improve the technology. The educational system needs to be connected with farmers and government to adjust research agendas and curriculum on water resources. Farmers and business people depend on the business environment and also contribute to shaping it for their own benefit. This explains the need of implementing communication mechanisms between leaders from the private sector, government, and academia to coordinate and collaborate on public policy and specific actions to improve water management. Engaging the private sector in water resource development is key to sustaining progress in the long term. Under this scenario, changes in national and local governments would less likely impact the continuity of policies and institutions that deal with water management. To achieve an effective collaboration, it is necessary to build an organizational structure that connects all actors and fosters collective activities. Organizations for water management at the basin level should be open institutions with the participation of all stakeholders.

1.08.3.3 Water Resource Management and the Regional Economic Strategy Water contributes with economic development through backward linkages that make the agricultural input supply sectors more dynamic, and via forward linkages strengthening agro-processing industries, trade, and other activities downstream in the value chain. Some of these activities are made feasible by the available technology for water agriculture. In addition, higher agronomic yields lower unit costs and impact on cheaper food available for the populace. Water also allows the cultivation of export crops that require a more regular supply and higher technology, and has an impact on off-farm employment. In a global scenario, countries and regions compete to offer the best location based on productivity to facilitate the value generating process that finally translates in prosperity. Globalization is putting pressure on countries and regions to

improve the determinants of competitiveness and they are adopting best practices in many aspects such as water management, water technology, agricultural technology, farm operation, infrastructure development, and human resource development, among other determinants of competitiveness. This competition leads to work on ways how to make a region distinctive from other competitor regions, to differentiate and achieve a distinct role in the country and in the world. Water then becomes an interesting resource to achieve regional differentiation, allowing competition and incentivizing innovation. Water management works as a channel to build a unique mix of strengths in the business environment, to differentiate and to improve the productivity potential of farms and agricultural businesses.

1.08.4 Water Resource Management, Institutions, and Implementation 1.08.4.1 Integrated Water Resource Management During the past couple of decades, agricultural water management has been approached within the integrated water resource management (IWRM) framework defined as a systematic process for sustainable development, allocation, and monitoring of water resource use in the context of social, economic, and environmental objectives (Cap-Net, 2009). This approach recognizes that water is used in various economic sectors, it has many uses, and as such it needs to be managed with an integrated approach. It recognizes that water management needs a multidisciplinary approach to deal with water sources and demands at the basin level. It also recognizes that water is a resource that needs to be sustained over time, and it is an instrument to achieve social, economic, and environmental goals at the basin level. Recently, this framework has been broadened to incorporate participatory decision making of all stakeholders. The IWRM is considered a paradigm shift that departs from traditional approaches in three ways: first, its multiple goals and objectives are cross-cutting so that this new approach departs from the traditional sectoral approach; second, the spatial focus is the river basin instead of single water courses; and third, there is a departure from narrow professional and political boundaries and perspectives and is broadened to incorporate participatory decision making of all stakeholders. The IWRM framework deals holistically with all water, all interests, all stakeholders, all levels and all relevant disciplines, and sustainable in all senses (Jaspers, 2001). In this framework, water is considered an economic good as well as a social good, and the market is given an important role for water pricing, and to deal with water scarcity, allocation efficiency, and environmental protection. Since the Dublin principles in 1992, there has been some progress in understanding the IWRM framework and its importance. However, there are many obstacles in its implementation process as a result of different sectoral interests, cultural values, and political constraints. Negotiation situates at the center of conciliating many thematic and special interests and beliefs at the time of allocating water resources. In addition, significant institutional and legal reforms are needed to reorient the traditional water management approach to a

Managing Agricultural Water

more integrated comprehensive management. Water development professionals still have difficulties in working together to come to a unique set of principles of IWRM.

1.08.4.2 Participatory Irrigation Management One of the most applied modalities in agricultural water management has been the participatory approach. Participatory irrigation management (PIM) refers to the involvement of irrigation users in all aspects and all levels of irrigation management. Users participate in all stages of the project life cycle; they begin with the participation on the design of the new (rehabilitation) project, and continue during its construction phase, including financing. Finally, and perhaps the most decisive stage, there is the participation of the users with the operation and maintenance of the system. It has been shown that PIM in real practice helps to solve some of the most prevalent problems with irrigation management, such as the inadequate water availability at the lowest outlets; poor condition/maintenance of the system; lack of measuring devices and control structures; inadequate allocation for O&M; inequitable distribution of water; and lack of incentives for saving water and poor drainage (Merrey, 2007). Since the beginning of the 1990s, this new perspective of irrigation management has been adopted in many developed and developing countries and has helped understand why some policies work better than others, and why some irrigation systems perform better producing higher impacts in terms of productivity and prosperity for the rural inhabitants. In many countries, PIM encouraged the creation of water user associations (WUAs) that took the responsibility of management of the irrigation systems once they were constructed. Taking management responsibilities demand from organizations the capacities to represent the users and institutions endowed with an adequate governance structure to take difficult decisions and solve the multiple obstacles, particularly those experienced during periods of water scarcity and conflict resolution. Before a WUA was formed, governments usually set a package of incentives for both users and irrigation agencies to make transfer programs sustainable. Irrigation agencies in many countries built processes of organization creation and strengthening, and delegated management of irrigation systems. One of the main objectives of the new management is to make the irrigation district financially sustainable, and to limit the amount of subsidies needed from the government. To finance the O&M activities, O&M fee schemes are usually implemented. The share in O&M contribution distributed between the government and users varies significantly among districts and countries; they are particularly biased toward a higher government contribution in those systems that were built with limited user participation. Sometimes the O&M fees are financed directly or indirectly through land or other type of agricultural taxes. To estimate the level of the fees, the portfolio of services of these organizations is identified, and costs are allocated. Once this information is available, usually the government and users enter into a negotiation process in which the responsibilities are assigned between the government and the WUA.

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User participation in defining service levels is crucial for achieving financial sustainability. The higher the user involvement in defining the set of activities that is needed, the more dynamic the future demand for those services will be and the higher the willingness to pay for the corresponding fee. In Mexico, for example, following the transfer of management of irrigation districts to WUAs, irrigation plans were prepared by WUA-hired managers, taking into account cropping plans, conveyance losses, and equitable distribution. These were then negotiated with the national irrigation agency to determine the final allocation, generally based on an arranged demand pattern. Operation and maintenance costs at the level of the secondary and below were met by user fees managed by the WUAs. In addition, these associations contribute with part of O&M costs (World Bank, 2009). In Mali, joint committees were established for O&M in every region. The committees had 5–10 representatives of producers and 5–10 representatives of the agency. These committees decided on types of services, costs including procurement matters, and water service fees. They also made decisions on the use of 50% of the user fees collected for O&M. In Chile, the national federation of WUAs was consulted in the design of an irrigation project. The federation and local WUAs played an active role in project preparation especially in discussions of service options and costs. Subsequently, the project incorporated the condition of WUA-approval for investment proposals and other project components. The fees usually have a component that covers some fixed costs and another variable component as a function of the water consumed or the irrigated area. Linking the fee to the irrigated area makes it easy to estimate it and to understand it. It is recommended that the easier to understand the fee structure, the simpler its administration. On the other hand, volumetric charges that are a function of the discharge generate the right incentive for water saving but are more difficult to implement because of the necessary water-measuring devices and control mechanisms that are costly and operationally complex. Agencies often favor a combination of the two methods. In Mexico, for instance, the agency charges the WUAs volumetrically at the turnout of the secondary canal, and the WUAs base their water charges to individual members on area irrigated and type of crop. In addition to unit area and volume, WUAs have also resorted to other bases such as charges for the entire season, which clearly favor the highvolume user. In some other irrigation districts, a property tax is employed as a charge based on the increase in the land values due to water availability. The higher the land valuation, the higher is the tax level as a consequence of available water supply. However, the problem observed with this system is that it does not generate the water-saving incentives that result with other mechanisms such as the volumetric approach, and usually these taxes are collected by municipalities, which generate discontinuity between the source and use of the resources. Another system, less frequently found, is the in-kind contribution system, mainly through labor, materials or both, applied directly to O&M activities. In Vietnam, the provincial and national governments finance schemes down to 150 ha for new irrigation development. Below this level, farmers must build the channels with the government providing support for

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survey and design, as well as contributing with materials, in some instances. After schemes are completed and taken over by farmers, each member must provide up to 20 person days per year toward maintenance of the tertiary as well as secondary systems (World Bank, 2009). But perhaps the most difficult hurdle to overcome is in replacing the assets of the irrigation districts. As any system, the infrastructure that is used for storing and delivering water to users depreciates. Buildings, canals, motors, pumps, vehicles, and heavy equipment need to be replaced once they complete their life cycle. If the government and/or the WUA do not implement a saving mechanism to cover these replacement costs, the sustainability of the system was compromised. Poor management leads to short service life with higher acceleration rates that then demand an increase in fees to cover future expenses. Provisions made for collecting these asset replacement funds by the WUA are usually not accepted by farmers who adopt a free-rider behavior and wait for government agencies to intervene when major depreciations do not let the system function properly. In parts of Vietnam, the estimate of costs of services included provision for depreciation of assets. This approach demanded knowing the location and conditions of all assets within the system, and to implement a training and preparation program on the formulation of an O&M program. Linking fees to services in the irrigation districts is important, particularly when an agency collects O&M fees and these funds are then allocated later to other government agencies, or when the budget of the O&M of the irrigation district is done independently of the collection effort. In these cases, there needs to be an agreement between the WUA and the government agency on mutual rights and responsibilities. The cost of collection by WUAs is lower relative to government agents and sanctions for nonpayment by individuals are easier when enforced by WUAs. The Philippines experience has shown that fee collection is better in systems where WUAs are organized and where they have a role and incentive for collection. In the pilot projects in Maharashtra, India, WUAs were allowed to retain a proportion of collections as a bonus (World Bank, 2009).

1.08.4.3 Lessons Learned from Participatory Management Since the 1990s, there has been an important effort on transferring the management of irrigation districts to WUAs. This new trend in water management has increased net irrigated area in large public irrigation projects. Important lessons have accumulated over time on various management schemes. Some lessons from the Philippines and Mexico and documented by the World Bank are presented below that exemplify the trend toward participation (these examples are taken from the Electronic Learning Guidebook for Participatory Irrigation Management).

1.08.4.3.1 The Philippines Just as the state’s involvement – or micro-management – was reaching its peak in the 1980s, there were countervailing forces appearing. In the Philippines, the process can be traced to the mid-1970s when President Marcos ordered the National Irrigation Administration to move toward self-financing. The

agency responded by withdrawing from the small communal systems which had once been self-managed, but had grown dependent on the government. This weaning process was accomplished through intensive grass roots organizing and capacity building both among farmers and within the agency itself. Water user associations were formed to take over operations and maintenance, and to contribute to capital costs of improvements. Beginning in 1980, this organizing approach was applied to the state-run systems that had no prior history of self-management. As with the management transfer in communal systems, the goal here was cost savings to the agency, both through direct recovery of water fees and the replacement of some low-level agency operational functions by association volunteers. Until recently, this modest level of joint management was the dream of irrigation policy reformers, and the Philippines served as the model. The paradigm was one of joint management where farmers would become management partners with the agency, and decisions would be made jointly. However, the relationship is asymmetrical; the state controls the technical expertise and subsidizes maintenance and improvements even in the canals operated by farmers.

1.08.4.3.2 Mexico Independently from the trend toward joint management at the lower ends of the system, a model of irrigation management transfer was evolving in Latin America, in response to structural adjustment pressures. This model constitutes the qualitatively different paradigm where the users dominate, and the state facilitates. In the mid-1980s when Mexico was in the throes of a debt crisis, the government was bankrupt. The large irrigation districts under Federal control suffered as maintenance was deferred and the productivity of unpaid, demoralized engineering staff declined. Out of necessity, the government reorganized the state irrigation agency to create the National Water Commission (or CNA in its Spanish acronym), with a mandate to turn over the management of the irrigation districts to associations of users created specially for this purpose. In 1990, Mexico transferred the first irrigation district to the users. By 1995, more than two-thirds of the country’s 3.2 million hectare network – divided into 80 irrigation districts – had been transferred to 316 irrigation associations. The transfer program was initially in the most productive irrigation districts, which were best organized and with the most commercially oriented farmers. The most important criterion for selecting districts was the potential of the user organization to become financially self-sufficient, with users paying the fees to cover the costs of operations, maintenance, and administration. What could the government offer the farmers as an incentive to accept higher costs for their irrigation? In fact, there was a carrot as well as a stick. The carrot was management autonomy. The farmers would be free to set their own rules for when to clean the canals, and how to distribute the water. The farmers would hire their own technical staff – engineers and accountants – to run their system. The canal would be theirs on a 20-year concession, which is in practice a transfer of ownership.

Managing Agricultural Water

However, there was also a stick. If farmers refused to take over management, the government could offer no assurance that the canal network could be kept in repair. The government in effect threatened to default on its conventional understanding with farmers regarding levels of subsidy in the irrigation sector because it no longer had the financial means to do so. The government, however, also promised and provided technical, organizational, and legal assistance in realizing the transfer. Many farmers, and particularly the commercially oriented ones, could not accept the risk that the irrigation infrastructure might collapse. They preferred to take over the management, and with a few exceptions, they have not looked back. They are paying much more for their water without the government subsidy, but the reliability and responsiveness of their new management structure is well worth the price. For them it is a win situation, and for the government as well. What are farmer’s comparative advantages when they are managing for themselves? They have direct incentives to manage irrigation water in a productive and sustainable manner; they offer an on-the-ground presence that even the most dedicated off-site agency staff cannot equal, and they have an intimate knowledge about their fellow irrigators. The state’s comparative advantage is in the depth of financial and technical resources and the regulatory and administrative capacity for managing water supplies to competing interests.

exercise of economic, political, and administrative authority to manage a country’s affairs at all levels, which comprises the mechanisms, processes, and institutions through which citizens and groups articulate their interests, exercise their legal rights, meet their obligations, and mediate their differences (United Nations Development Programme, 2001). Rogers (2002) argued that governance is not restricted to the perspective of government as the main decision-making political entity, and governance covers the manner in which allocative and regulatory politics are exercised in the management of resources (natural, economic, and social) and broadly embraces the formal and informal institutions by which authority is exercised. According to Rogers (2002), there are different models of water governance:

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• 1.08.4.4 Institutions and Water Governance Before presenting the argument of the need of institutional and organizational reform, some key concepts need to be defined. According to North (1990), institutions are the rules of the game in society. They are social arrangements that shape and regulate human behavior and have some degree of permanency and purpose transcending individual human lives and intentions. In agricultural water management, these types of institutions are user associations, rules of water allocation, market mechanisms, and property rights. Organizations refer to the group of people with shared goals and formally defined roles, such as water associations, government irrigation agencies, water companies, nongovernmental organizations, and regulation bodies. A policy is ‘‘a set of interrelated decisions taken by a political actor or group concerning the selection of goals and the means of achieving them within a specified situation where these decisions should, in principle, be within the power of those actors to achieve’’ (Howlett and Ramesh, 1995; Jenkins, 1978). On the other hand, governance is the way authority is organized and executed in society, and often includes the normative notion of the necessity of good control. The Global Water Partnership defines water governance as ‘‘the range of political, social, economic, and administrative systems that are in place to develop and manage water resources, and the delivery of water services, at different level of society’’ (Rogers and Hall, 2003). Governance is therefore a broad term that includes institutions, organizations, and policies. The World Bank broadens the definition to include the process by which those in authority are selected, monitored, and replaced, and the effectiveness of government in implementing sound policies (Jayal, 1997). The United Nations defines it as the

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The bureaucratic politics and process model. This model is based on political–bureaucratic bargaining in a federal system. Its focus is typically on the executive branch, with the elected legislature hardly in the picture. The congressional behavior model. A second federal model concentrates on the elected congress, with the view that to understand congressional behavior is to understand that congressmen are single-minded seekers of reelection. It follows from this that congressmen’s goals are to improve the welfare of their constituents in the shortest possible time frame. The interest group model. In some cases, legislators see only a few dominant interests involved with water policy. These interest groups often have overlapping concerns and overlapping memberships. It is also valuable to carry out an interest group analysis of the feasibility of pursuing specific governance goals. They examined the likelihood that the various interest groups would be powerful enough to influence the investment and management decisions in their direction. Principal–agent theory. Principal–agent models have been employed in many different academic fields, including economics, in order to explain relationships among actors in which the consumer is the principal and the producer is the agent; and in various political science subfields, in which members of the legislatures are the agents of their constituents, or bureaucrats are agents of the executive, or the governments of Third World countries are the agents of international lending institutions, and so on. Regime theory and public choice. The use of the concentration or diffusion of costs and benefits of public choices to predict what decision-making system will prevail.

It has been found that water governance for agriculture is important for the management of irrigation systems and rainfed agriculture involving different stakeholders to achieve more effective water allocation. Good governance in irrigation systems matters for achieving high levels of economic, social, and environmental performance. There are some necessary conditions for good governance: inclusiveness, accountability, participation, transparency, predictability, and responsiveness. A failure on these conditions leads to poor governance and difficulty to deal with system problems and presents a risk in terms of sustainability. Better governance will lead to higher levels of prosperity of the users

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of the agricultural systems and becomes an essential input of country and regional development. There is also a strong relationship between the level of governance of water resource systems and the income per capita, lower infant mortality, and higher literacy (Kaufmann, 2005). New institutions and governance are needed to implement water management if the goals are to achieve food security, economic growth, and poverty reduction under favorable environmental conditions. This demands institutions and organizations responsible for good agricultural water management and policymaking. There is a need to reform not only the organizations at the national and international level that deal with water management, such as the irrigation agencies and multilateral institutions, but also local and informal institutions that act at the basin level that need to be redesigned and in some cases built from scratch. The current institutional arrangement has failed to meet the challenge to respond to new technologies and rules to deal with the challenging goals of developing agriculture to reduce poverty, and generate prosperity with equitable growth under minimum environmental impacts. In the water sector, as with any sector in the economy, some institutional arrangements have shown little promise so far. On the one hand, agricultural markets in developed and developing countries present high distortions that generate strong market failures that need government intervention. On the other hand, the government itself frequently fails to make the right adjustments to align market outcomes. In the irrigation sector, arrangements for authorization, payment, and accountability impact on service provision (Huppert et al., 2001). The failure of public organizations to offer socially and environmentally optimal allocations of water resources has induced the creation of new institutional schemes. The decades of the 1990s and the 2000s have experienced an upsurge of private companies and market mechanisms not only in the construction of water systems for agriculture but also for their operation and maintenance. Some of these new organizations present a heavy involvement of the private sector. Privatization of O&M has been a component of many irrigation programs, particularly where the operation of these systems required specialized skills that individual farmers did not have. Two major reforms have been put in place in many countries to generate economic incentives for water allocation: water pricing and tradable water rights. Water pricing mechanisms are used to create incentives for water conservation but require volumetric measurement devices that are expensive and difficult to control. In addition, the strong legacy of government subsidies and public intervention in water management has generated famers’ resistance to pay for irrigation service. In addition, investment cost recovery demands such high fees that unless there exist financial markets it is difficult to obtain payments from farmers. The second privatization mechanism that has been used frequently in many countries during the past two decades is the tradable water rights. The idea behind the water rights market is that the user who has a higher value use of the resource would be willing to pay more for the right to use the water. In other words, the market allows water to move from lower to higher value uses. However, these markets are

complicated to build since specific topologies of irrigation systems to transfer water from buyer to seller are required. They also need institutional arrangements to control, facilitate, and enforce transactions, and protect against negative impacts on third parties when water is transferred (Easter et al., 1998; Rosegrant and Binswanger, 1994).

1.08.4.4.1 Water management principles In 1992, at the International Conference on Water and the Environment, convened in Dublin, Ireland, four main principles of water were adopted that have since then shaped water management. The first principle is that water is a finite and vulnerable resource, essential to sustain life, development, and the environment. This principle recognizes that water is critical to sustaining life, and it is a finite resource because of the economic and technical complications of transferring water from one place to another. This principle also suggests a holistic approach to deal with water management in which all dimensions, social, economic, and environmental are taken into account. The second principle establishes that water development and management should be based on a participatory approach, involving users, planners, and policymakers at all levels. It suggests democratization in the decision-making process when allocating water among uses, and the importance of the input of multiple stakeholders in the design, construction, and operation and maintenance of systems such as irrigation districts. The third principle is that women play a central part in the provision, management, and safeguarding of water. In many parts of the world, women are the ones who collect and safeguard water for domestic and agricultural use. However, women do not have as important a role as men in water management, and their involvement has been shown to be important for achieving higher levels of efficiency and sustainability. The fourth principle, perhaps the most controversial at that time, and the one that has created a major impact in water management, is that water has an economic value in all its competing uses and should be recognized as an economic good as well as a social good. This principle highlights the need of assigning value to water not only as a social good but also as an economic good if water is to be allocated efficiently and if water systems need to be sustainable in all dimensions. This principle has generated debate for its implicit recommendation of charging for water.

1.08.5 Water Management and the Environment Almost by definition, water management impacts the environment; changes in agriculture cause modifications in land cover and watercourses, and degrade ecosystems and impact their services for human development. In order to increase productivity, land and water are manipulated in different ways with consequences on the environment. Some of these manipulations include the following (Falkenmark et al., 2007). (1) Shifting the distribution of plants and animal, when the native vegetation is cleared and is replaced by crops and wild animals are replaced with livestock. (2) Coping with climate

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variability to secure water for crops. As water is a key material for photosynthesis, crop productivity depends intimately on securing water to ensure growth. (3) Maintaining soil fertility. The conventional way to secure enough air in the root zone is by drainage and ditching through plowing to ensure that rain water can infiltrate, a process that leads to erosion and the removal of fertile soil by strong winds and heavy rain. (4) Coping with crop nutrient needs. The nutrient supply of agricultural soils is often replenished through the application of manure or chemical fertilizers. (5) Maintaining landscape– scale interactions. When natural ecosystems are converted to agricultural systems, some ecological processes (such as species mobility and subsurface water flows) that connect parts of the landscape can be interrupted. This can have implications for agricultural systems as it can affect pest cycles, pollination, nutrient cycling, and water logging and salinization (Lansing, 1991; Cumming and Spiesman, 2006; Anderies, 2006). The Millennium Ecosystem Assessment, an international assessment by more than 1300 scientists of the state of the world’s ecosystems and their capacity to support human wellbeing, identified agricultural expansion and management as major drivers of ecosystem loss and degradation and the consequent decline in many ecosystem services and human well-being (MEA, 2009). This study showed that by year 2000, a quarter of the global land cover had been converted for cultivation, with cropland covering more than 50% of the land area in many river basins in Europe and India and more than 30% in the Americas, Europe, and Asia. It also showed that the development of water infrastructure and the regulation of rivers for many purposes, including agricultural production, fragmented rivers and generated impoundment of large amounts of water (Revenga et al., 2000; Vo¨ro¨smarty et al., 2005). Failure of agricultural water management increases the effect of natural and human-induced disasters, such as droughts and famine, on poor people. The rural poor, who are highly vulnerable, have a high dependency on the ecosystems where they farm or live and adequate water management diminishes the negative impact of low precipitation events (Silvius et al., 2000; WRI et al., 2005; Zwarts et al., 2006). Issues related to gender in management are also important in managing water for agriculture, as well as recognizing the importance of extreme events for the evolution of ecosystems. Water use for irrigation for agricultural production alters the excess supply of water for other uses, such as surface and groundwater for aquatic use downstream. The spatial water distribution changes significantly after an irrigation system is built and reduces the availability of water for downstream uses, impacting ecosystems. Irrigation infrastructure, dams, and canals alter the waterscape and impact the ecosystem, and irrigation water alters also the hydrologic cycle at the basin level. Part of the transfer water from streams or aquifers moisturizes soil and infiltrates, and the rest either is used by the plants or is evaporated. Unfortunately, increasing agronomic yields require not only water use but also the application and increased use of agrochemicals. Irrigation indirectly contributes to the application of chemicals and creates a negative impact on water quality with externalities downstream where water is also needed. The challenge is then how to increase agricultural

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production without compromising the sustainability of ecosystems needed by society in general. General recommendations are being contemplated to deal with the negative impact of water use in agriculture (Falkenmark et al., 2007). The implementation of an integrated approach to manage land and water resources and ecosystems that acknowledge the multifunctionality of agroecosystems in supporting food production and ecosystem resilience are recommended. To accomplish this, a better understanding of the services that are generated by agroecosystems and the value of biodiversity is required. Unfortunately, zero-environmental-cost agricultural water management is almost unobtainable. This obligates to understand the value of maintaining biodiversity and the interaction of ecosystems and their future water requirement to sustain ecosystem health and biodiversity and to evaluate how these benefits compare to short-term agricultural development. Given that new irrigation systems are converting land from its original ecosystems and regulating rivers are creating an impact on spatial water distribution, it is recommended to improve existing agricultural systems to obtain an increase in production as opposed to the expansion of agriculture. There is significant room for improvement in terms of technologies and management practices with techniques more environmentally safe to lower the impacts from the additional agricultural production needed to satisfy increasing food demand. There is also a need to reduce or reverse the ecosystem degradation through rehabilitation and restoration of agricultural systems. Mitigation of negative impacts on ecosystem services is achieved through an integrated management of land, water, and other dimensions of ecosystems at the watershed level. This approach requires a participatory process in which users can understand the benefits and costs of various development options, and to foster stakeholder discussion about the tradeoffs. Tools are available to assist this process, from economic evaluation to environmental assessment techniques, and should be complemented with more sociologic and human interaction methodologies. New tools need to be developed to monitor the evolution of ecosystems and to generate the required feedback to take the appropriate corrective decisions. Unfortunately, the agricultural development process and its impact on the environment are full of uncertainties that need to be dealt with when evaluating trade-offs and taking decisions. Using existing and developing new techniques to deal with uncertainty within the planning process is necessary. Tools such as scenario planning could improve the assessment and learning of the development decisions and their potential impacts on the environment. Increasing food demand requirements and minimizing its negative impact on the environment require building new institutions and organizations that could better address the different development options and their impacts on the ecological functions. To generate the best setting for decision making, it is necessary to educate all stakeholders about the ecological services of ecosystems, and the social and economic benefits of alternative policy options. A higher level of abstraction is needed so that interested people from different disciplines can internalize the relevant information for decision making.

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1.08.6 Water for Agriculture and Poverty Reduction High-income countries have shown important increases in agricultural productivity and lower-income countries have benefited less from improved varieties from the Green Revolution, which has been a major contributor to growth in staple crop production in the developing world. Production growth in sub-Saharan Africa was based almost exclusively on the extension of cultivated area, and in Asia, irrigation was part of the explanation since most of the innovations in agricultural varieties demanded supplemented water via irrigation. In developing countries, the population will remain predominantly rural until 2020 when the size of the rural population declines due to slower population growth and rapid urbanization (World Bank, 2008). Poverty is more prevalent in the rural areas and the World Bank (2008) estimates that gross domestic product (GDP) growth generated in agriculture is, on average, 4 times more effective in alleviating poverty and benefiting half of the poorest population than growth generated outside agriculture. This effect declines as the countries have higher incomes. There is a clear relationship between water management and poverty reduction. Case studies from research and international institutions show that water management is a good investment that not only contributes to poverty reduction, but also is generally cost efficient, and has the potential to generate wealth. As a result of this linkage, the global community has united to fight poverty through actions that bring different interests and organizations together in effective partnerships around the MDG agenda. Following the conceptual framework developed by the Poverty–Environment Partnership (2002), water management contributes to poverty reduction in four dimensions:

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Enhanced livelihood security. Water is considered an input to livelihood activities and a determinant of health and productivity of ecosystems. Water in agriculture positively impacts the ability of poor people to use their agricultural land and capabilities to make a living in conditions of greater security and sustainability. Ensuring continuity in water flows and minimum levels of water quality is essential for maintaining the integrity of ecosystems. Making sure that adequate and reliable water supplies are available for agricultural activities (including livestock, aquaculture, horticulture, and other types of production) is key to poverty reduction throughout the developing world. Therefore, the adequate design of water irrigation districts makes water available for activities that impact directly the livelihood of rural habitants. Reduced health risks. Water-borne and water-related vectorborne diseases, such as diarrhea and malaria, are the main killers in many parts of the developing world, and, in particular, affect children and other vulnerable groups. Some of these diseases have their origin in irrigation systems and adversely affect the most vulnerable population, women, and children, generating disabilities, poor nutritional conditions, and, eventually, death. Well-designed irrigation systems, with the right hydraulic infrastructure and with an adequate administration and operation, together with educational campaigns, is the most effective

strategy to improve health in the rural areas. It has been found that, in many cases, the economic benefit from these ameliorating activities is higher than the benefit derived from an increase in the regional economic value-added in many parts of the developing world. Integrating water management in agriculture with health systems development is one of the most effective strategies to reduce poverty through health improvement of the rural population. Every day diarrheal diseases cause nearly 5000 deaths, mostly among children under 5 years of age (United Nations, 2006). The World Health Organization estimates that the number of deaths from infectious diarrheas amounts to 1.8 million for all age groups, with a heavy toll among children under 5–1.6 million deaths. Malaria kills about 1 million people in the world every year, mainly in Africa, and about 80% of these cases are children under 5 years of age. It is estimated too that 160 million people are infected with schistosomes and 133 million suffer from high-intensity intestinal infections. These diseases have a direct impact on the welfare of the population and a long-term effect on future generations.

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Reduced vulnerability. Water in agriculture is often too much, too little, or too contaminated. It is rare to observe equilibrium of natural water supply with the right amount of water potentially demanded by the agricultural systems. This variability of water supply generates environmental, economic, and social threats as a result of sudden impact shocks and long-term trends. Droughts, floods, and highintensity rainfall have the potential to destroy livelihoods and increase poverty in the rural areas. In addition, water management in agricultural systems should compensate for the negative impacts of climate change, pollution, and solid degradation. Well-designed and managed agricultural systems reduce the resilience of the poor and help stabilize the income variability of rural landholders, and participants of the agricultural value chain. Pro-poor economic growth. Agricultural water management has the potential to generate the economic growth needed to reduce poverty in the rural sector and, in particular, to create income-generating opportunities for the poor. Water supply is a necessary input in productive activities and allows technologies adoption. Agricultural water management opens the opportunity for entrepreneurs to participate in ventures upstream and downstream of the agricultural value chain.

The potential economic dynamism generated by the induced economic activity by an improved water management will generate returns and benefits to the local economy with many multiplier effects. National and regional economies benefit with significant improvement in the management of agricultural systems, especially if these activities are complemented with measures to make input markets work, additional diversification opportunities, and investment in related infrastructure. Water management improvement should be accompanied by institutional strengthening and knowledge-generation initiatives to increase its impact on poverty reduction. Based on the previous framework, there are evident opportunities to reduce poverty through water management in

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agriculture. The UN Task Force on the MDG found four core areas on which it is necessary to work to eliminate the constraints (United Nations, 2006). First, it is necessary to work on policy, legal and regulatory reform, including issues of rights of access to water. Attention should be given to work on policies that target needs and opportunities of the poor for improved access to water for agriculture. Second, it is necessary to plan and select technology choices that are consistent with poverty reduction targets, assessing the possible impacts on the most vulnerable people and the water resources. Third, it is necessary to develop financial mechanisms such as investment incentives and cost-recovery mechanisms to have credit and financial management systems and to create a regulatory regime and climate where private investment is encouraged. Finally, institutional reform and coordination among government agencies are needed to deal with agricultural water management to support the investment and involvement from the private sector and other regional actors, and improve the management of existing systems.

1.08.7 Water Management of Rainfed Agriculture Water productivity in rainfed agriculture is low despite being the traditional practice on 80% of the world’s agricultural area, which presents an opportunity for investment to improve productivity and boost agricultural yields. In many communities in developing countries, rainfed agriculture remains as the primary source of food, particularly grain. Historically, many developing countries that practiced rainfed agriculture had chosen to increase production through brute force by simply expanding their agricultural area. Despite yields having increased by a significant amount in these countries, they still remain far from the yields found in developed countries such as the United States and Europe, which serves to demonstrate the potential for yield growth in developing areas (Rockstro¨m, 2007). On average, these developing countries have only reached approximately 30% of their achievable yields, while some countries such as Yemen and Pakistan only reaching approximately 10% of their achievable yield. Furthermore, historical evidence has demonstrated a ‘‘growing yield gap between farmers’ practices and farming systems that benefit from management advances’’ (Wani et al., 2003), which serves to show that neglecting water management practices not only prevents farms from achieving higher yields, but in fact it diminishes their yield as well. Moreover, it has been shown that improving rainfed agriculture has led to decreased poverty rates (Irz and Roe, 2000), which further emphasizes the need for investment in rainfed agriculture to heighten productivity. The key problem facing farmers that has hindered yield growth cannot simply be generalized as the amount of rainfall: depending on where in the rainfall zone spectrum an area lies, the core limiting factor varies. In arid regions, the amount of rainfall does in fact prove to be the largest problem for rainfed agriculture. However, in other regions, such as the semiarid and dry subhumid zones where the minimum crop water need is already met, the problem of yield growth becomes focused on high variability of rainfall. Challenging previous beliefs that the amount of rainfall was the main

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limiting factor for higher yields, this sporadic availability of water has helped to frame rainfed agriculture by demonstrating that basic water management strategies can double yields, on average. This variability in rainfall, characterized by seldom and high-intensity rainfall as well as frequent dry spells and high-intensity droughts, creates an insufficient water supply to meet water demand, otherwise known as water stress. Water stress has two primary divisions: one is the dichotomy between dry spells, which are 2–4-week periods during the rainy season that temporarily halts production during critical stages of growth, and droughts, which are a less common (once a decade) lack of rainfall that entirely halts crop growth, while the second division is one between man-made variability (agricultural) and natural variability (meteorological). Further exacerbating the problem of water stress is the variability in amount of rain that reaches the crops’ roots through soil moisture, which is generally only 70–80% of total rainfall, and sometimes can be as low as 40% in areas of poorly managed land. These problems incited by poor water management techniques are referred to as agricultural dry spells and droughts, and provide additional evidence that investment in water management can drastically improve yields. When rainfall is not efficiently used for plant absorption due to faulty management practices such as poor soil fertility or land degradation, the shortage of food is blamed on an agricultural drought. For example, in semiarid regions, nonplant growth rainfall, which includes drainage, nonproductive evaporation, and runoff, account for up to 85% of rainfall. Transpiration, which directly attributes to plant growth, only accounts for up to 30% of the rainfall. In certain areas with severely degraded land, only 5% of rainfall is used productively for plant growth. Rainfed agriculture yields are often limited by deficient soil fertility that comes as a result of nutrient depletion, loss of organic matter, and other forms of soil degradation. Studies in India have shown that this phenomenon is oftentimes human induced, where practices such as subsistence farming have depleted the soil of many of the necessary plant growth nutrients. Fortunately, problems concerning soil fertility can be easily rectified through investment in water management, such as injecting micronutrients into the deficient soil, which has been shown to substantially increase crop yields and rainwater productivity in India. As a result of this increased production, economic returns also increased by 50–75%. These studies further emphasize the capabilities and the need for investment in rainfed systems. The solution to the low productivity of rainfed systems is much easier said than done. A multitude of challenges stand in the way of simply investing more in rainfed agriculture and increased yields, the most fundamental of which is the adaptation and implementation of new agricultural innovations to every farm worldwide. Oftentimes, what drive innovation to be realized and adopted are social and ecological crises. However, simply waiting until the advent of the next crisis is not a viable option. Moreover, most farming households are small and marginalized and do not necessarily have access to the latest agricultural innovation. Rainfed areas also tend not to have the necessary infrastructure to implement

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innovations because throughout history, only high-potential irrigated areas have received large innovations, leaving these smaller areas in the dust. Other roadblocks include ‘‘limited information of the options available, social and economic constraints to adoption, lack of enabling environments and backup services, poor market linkages, and weak infrastructure’’ (Rockstro¨m, 2007). Developing avenues for better water management practices, such as through governance, policy, institutions, practices, and technologies, requires more attention than what has currently been devoted to it. Current strategies aim to make more efficient use of rainfall by increasing plant water availability and plant water uptake capacity through external and in-site water harvesting systems, soil and water conservation, evaporation management, and integrated soil, crop, and water management. While there are efforts to reduce water loss by increasing canopy cover, most strategies focus on securing more water produced by rainfall. In order for these strategies to be successfully implemented, they require a larger focus on small-scale water harvesting. Since small households make up a large proportion of farmers, unless these techniques are financially and structurally feasible, they simply cannot be used. The structure of rainwater harvesting is divided into three parts: ‘‘a watershed area used to produce runoff, a storage facility (soil profile, surface reservoirs, or groundwater aquifers), and a target area for beneficial use of the water (agriculture, domestic, or industry)’’ (Siegert, 1994). Another angle of attack that accounts for rainfall variability includes maintaining yields throughout dry spells and droughts. One important strategy lacking prevalence is the technique of supplemental irrigation, which collects runoff from external rainwater harvesting systems and transfers it to rainfed cropland. This technique, when compared with strictly rainfed systems, has shown substantially higher crop yields, with increases in yield ranging from approximately 30% to 400% (Oweis and Taimeh, 1996). Arguably, the most valuable quality of supplemental irrigation is its ability to supply water to a region when it is undergoing a dry spell, thus diminishing the frequency of interruptions to crop cycles. Studies have shown that 50–200 mm of supplemental irrigation is enough to combat the yield-reducing effects of dry spells (Oweis and Taimeh, 1996). Furthermore, the cost of supplemental irrigation systems is low, even for family-sized farms and small communities, in that the cost of a reduced yield due to dry spells is effectively eliminated. The technique of supplemental irrigation has already been adopted and implemented by a multitude of commercial farms in countries such as Australia, South Africa, and India; however, the benefits of such a system must be realized through simultaneous practice of other water-management techniques in order to maximize crop yield.

1.08.8 Policy Actions for the Future IWMI (2007) has completed an assessment of water management for agriculture and recommended eight policy areas, recognizing first that different actions are required for different situations, that actions for the sub-Saharan Africa will be

needed in terms of infrastructure, and where infrastructure has already being developed, as in Asia, the improvement of water availability relied on increases in productivity, relocating supplies, and rehabilitating systems. The eight policy actions are discussed below:

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Change the way we think about water and agriculture. To achieve food security, reduce poverty, and conserve the ecosystems, it is necessary to think differently about water. A broad focus should be adopted, with reformed institutions and viewing rain as the ultimate source of water that should be managed. Agriculture should be seen as a system integrated with other uses, providing services and integrating with other ecosystems. Fight poverty by improving access to agricultural water and its use. To effectively fight poverty, it is necessary to secure water access to smallholder farmers, using pro-poor technologies, and investing in roads and markets. In addition, systems that have multiple uses, such as aquaculture and agroforestry, are an ideal solution to alleviate poverty. Good governance is fundamental to any poverty-reduction strategy. Good governance means creating a fair legal, policy, and regulatory framework in which the rights of people to access water resources for agriculture are secured. It deals also with improving the effectiveness, accountability, and transparency of government agencies, ensuring the participation of the poor in decision making and enhancing the role of civil society guarantying basic security and political freedom (United Nations, 2006). Manage agriculture to enhance ecosystem services. Agricultural management could enhance other ecosystems to promote services beyond the production of traditional agricultural products. Water and land use will probably be intensified in the future but should be articulated with other services. Increase the productivity of water. More food per cubic meter of water should be the goal to increase effective water supply and reduce effective water demand. It is estimated that 35% increase in water productivity could reduce additional crop water consumption up to 80%. The poor can benefit from water productivity gains as well as largeholder farmers by introducing higher-value products. Upgrade rainfed systems – a little water can go a long way. The way to upgrade rainfed systems is by conserving soil moisture and providing supplemental irrigation where feasible. This action has a great potential to increase the income of an important number of farmers, especially in water-deficit regions such as sub-Saharan Africa and parts of Asia. Mixed crop and livestock systems have a good potential for improving the productivity of these systems. Adapt yesterday’s irrigation to tomorrow’s needs. Emphasis should be given to adapt current irrigation systems to future demands. Rehabilitation, modernization, and management techniques should be used to increase water productivity and its impact on poverty alleviation. Reform the reform process – targeting state institutions. Government institutions should be reformed if new management approaches are to be adopted and implemented. A collaborative process of all actors, private sector, and other institutions, is important to ignite a renovated process to satisfy local needs.

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Deal with trade-offs and make difficult choices. There should be good information to allocate water among conflicting uses. Informed multi-stakeholders are necessary to achieve effective negotiations and to make decisions about the use and allocation of water. Reconciling competing demands on water requires transparent sharing of information.

1.08.9 Summary Agricultural water management deals with the administration of a key input to agricultural production, and up to now it has been responsible for the majority of water use in the world, has improved nutrition, alleviated poverty, and increased production in 2.5 times the level in four decades. Agricultural water management has had the simultaneous challenge of meeting future food needs, making farms profitable and reducing poverty. New technologies are needed to intensify land and water use and to increase agronomic yields to produce the additional food to feed the world population and minimizing the impact on the environment. However, most importantly, agricultural water management needs to provide efficient solutions to improve the income levels of the farmers and rural people. These water resource challenges require better allocation of the resource using the social and economic value of the resource as distribution criteria. Agriculture has induced the degradation of many ecosystems, including those that are essential for food production. Aquifer contamination and depletion, drainage of wetlands, surface water contamination, land degradation and erosion, and aquatic ecosystems conflict are just some examples of these impacts. Therefore, water in agriculture should be managed by taking into account the array of services of the ecosystem, which are crucial for society. Adequate planning needs to take into account the trade-off between water for food production and ecosystem services, and address the social impact of poor rural people who often suffer the consequences of environmental negative externalities. Agricultural water management has the potential to become the best road to improve agricultural practices through technology adoption and diminish the environmental impact of the use of water resources. Water managers for agriculture will face many challenges while trying to accomplish multiple objectives. Sacrifices will be needed and so the appropriate institutional arrangement is to deal with these trade-offs. That is why an integrated approach to land use, water allocation, and ecosystem sustainability will be essential. In addition, understanding the tradeoffs between the conflicting goals will require an increase in the body of available knowledge to have a more accurate approximation as to the effect of irrigation to the environment and the effect of climate change on the future water supply for agriculture. The trend of irrigation investment has decreased by half during the past two decades compared to the growth experienced in the 1960s and 1970s. Constraints are larger as easy sites are already been exploited, groundwater sources are unsustainable in many places and the world is now more concerned with environmental and social issues that set a higher

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standard in agricultural water management, a situation that generates the incentive for new technology development. At the same time, nation and regional development processes have generated new uses for water that are competing with agricultural use. One of the great setbacks of agricultural water management is the severe failure of many institutional arrangements at the national and local basin levels to deal with water management. New rules and incentive mechanisms need to be put in place in order to face the future challenge to reflect our understanding of the interaction of water resources with ecosystems and human activity. Irrigation systems involve people with different interests, and therefore good governance is needed as a framework (institutional and administrative) to allow cooperation, coordination, and, most importantly, conflict resolution. Finally, water management for agriculture could become an interesting tool to achieve regional differentiation, allowing competition and incentivizing innovation. Water management could help building the strengths needed to create a differentiated business environment and therefore creating the conditions to achieving higher levels of regional and national prosperity through increments in productivity.

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Ramirez-Vallejo J and Rogers P (2004) Virtual water flows and trade liberalization. Journal of Water Science and Technology 49(7): 25--32. Ramirez-Vallejo J and Rogers P (2009) Failure of the virtual water argument: Possible explanations using the case study of Mexico and NAFTA. In: Biswas AK and Cline SA (eds.) Global Change: Implications for Water and Food Security. Washington: IFPRI. Revenga C, Brunner J, Henniger N, Kassem K, and Payner R (2000) Pilot Analysis of Global Ecosystems, Freshwater Systems. Washington, DC: World Resources Institute. Rockstro¨m J (2007) Water in rainfed agriculture. In: Molden D (ed.) Water for Food, Water for Life. A Comprehensive Assessment of Water Management in Agriculture, ch. 8. London: Earthscan, and Colombo: International Water Management Institute. Rogers P (2002) Water Governance in Latin America and the Caribbean. Washington, DC: Inter-American Development Bank, Sustainable Development Department, Environment Division. Rogers P and Hall A (2003) Effective Water Governance, GWP Technical Committee Background Paper 7. Stockholm: Global Water Partnership. Rosegrant M, Cai X, Cline S, and Nakagawa N (2001) The role of rainfed agriculture in the future of global food production (invited background research paper). In: International Freshwater Conference. Bonn, Germany. Rosegrant MW, Agcaoli M, and Perez ND (1995) Global Food Projections. Canada: Renouf Publishing. Rosegrant MW and Binswanger H (1994) Markets in tradable water rights: Potential for efficiency gains in developing country water resource allocation. World Development 22(11): 1--11. Rosegrant MW and Cai X (2002) Global water demand and supply projections: Part 2. Results and prospects to 2025. Water International 27(2): 170--182. Rosegrant MW and Cai X (2004) Implications of water development for food security. In: Lawford R, Fort D, Hartmann H, and Eden S (eds) Water Science, Policy, and Management,. Water Resources Monograph Series, vol. 16, 422pp. Washington, DC: American Geophysical Union. Seckler D, Molden D, and Sakthivadivel R (2003) The concept of efficiency in waterresources management and policy. In: Kijne JW, Barker R, and Molden DJ (eds.) Water Productivity in Agriculture: Limits and Opportunities for Improvement, Comprehensive Assessment of Water Management in Agriculture Series No. 1, pp. 37–51. Wallingford and Cambridge, MA: CABI Publishing. Siegert K (1994) Introduction to water harvesting: Some basic principles for planning, design and monitoring. In: Water Harvesting for Improved Agricultural Production, Proceedings of the FAO Expert Consultation, Water Report 3. Cairo, 21–25 November 1993. Rome: Food and Agriculture Organization. Silvius MJ, Oneka M, and Verhagen A (2000) Wetlands: Lifeline for people at the edge. Physical Chemistry of the Earth B 25(7–8): 645--652. Travis WP (1964) The Theory of Trade and Production. Cambridge: Harvard University Press. United Nations (2006) Millennium Development Goals Report. New York: United Nations. United Nations (2008) Trends in Sustainable Development. Agriculture, Rural Development, Land, Desertification and Drought. New York: United Nations United Nations Development Programme (2001) Partnerships to Fight Poverty. Annual Report. New York. United Nations Development Programme (2001) Partnerships to Fight Poverty. Annual Report. New York. Vo¨ro¨smarty CJ, Le´veˆque C, and Revenga C (2005) Fresh water. In: Hassan R, Scholes R, and Ash N (eds.) Ecosystems and Human Well-Being: Current State and Trends – Findings of the Condition and Trends Working Group. Washington, DC: Island Press. Wallace JS and Batchelor CH (1997) Managing water resources for crop production. Philosophical Transactions of the Royal Society London B 352: 937--947. Wana J, Griffithsa R, Yinga J, McCourtb P, and Huanga Y (2009) Development of drought-tolerant canola (Brassica napus L.) through genetic modulation of ABAmediated stomatal responses. Crop Science Society of America 49: 1539--1554. Wani SP, Pathak P, Sreedevi TK, Singh HP, and Singh P (2003) Efficient management of rainwater for increased crop productivity and groundwater recharge in Asia. In: Kijne JW, Barker R, and Molden DJ (eds.) Water Productivity in Agriculture: Limits and Opportunities for Improvement, Comprehensive Assessment of Water Management in Agriculture Series No. 1, pp. 199–215. Wallingford and Cambridge, MA: CABI Publishing. Wichelns D (1999) Economic efficiency and irrigation water policy with an example from Egypt. Water Resources Development 15(4): 543--560. Winpenny J (2003) Report of the World Panel on Financing Water Infrastructure. World Water Council. World Bank (2003) Managing Water as an Economic Good: Rules for Reformers, Water Resources Sector Strategy. Washington, DC: World Bank.

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1.09 Implementation of Ambiguous Water-Quality Policies DH Moreau, University of North Carolina, Chapel Hill, NC, USA & 2011 Elsevier B.V. All rights reserved.

1.09.1 1.09.2 1.09.2.1 1.09.2.2 1.09.3 1.09.3.1 1.09.3.2 1.09.4 References

Nonpoint Sources and the CWA Intrastate Cases Neuse River Nutrient Management Strategy Jordan Lake Stormwater Rules Interstate Nonpoint Management Mississippi River Basin and Hypoxia in the Gulf of Mexico Chesapeake Bay Program Summary and Conclusions

Ambiguities in water policies may create significant barriers to implementation and lead to unpredictable outcomes, especially in the United States’ federal system where the national government has made some parts of water policy unambiguous but left other parts to states with ambiguous mandates. The Clean Water Act (CWA) is a case in point. The goal of the act is to protect and enhance the quality of all waters of the United States to levels that are sufficient to support their state-designated uses, at a minimum to make them fishable and swimmable. Policies set forth in the Act establish several levels of technology-based effluent limits for point sources, publicly owned wastewater treatment plants, and a large number of categories of industrial dischargers. Enforceable permits issued to those sources contain effluent limits, monitoring protocols, and reporting requirements. The permit program also covers other sources such as urban stormwater runoff and concentrated animal feeding operations (CAFOs). There is little ambiguity if the least stringent effluent limits are sufficient to satisfy water-quality standards, but several ambiguities arise when minimal requirements covering these sources are not sufficient to satisfy water-quality standards. These ambiguities are especially important when sources of degradation are dominated by agriculture and urban stormwater runoff. In these cases, states are obligated to establish maximum allowable loads for those segments of water bodies that have been designated as being degraded. Then, states are obligated to allocate a portion of the allowable load to point sources, another portion to nonpoint sources, and reserve a third portion as a factor of safety. If more stringent effluent limits on point sources are insufficient to upgrade water quality to satisfy standards, then states are required to submit plans to control nonpoint sources to the maximum extent practicable (MEP). However, most agricultural sources are excluded from federal permits. Management of those sources is largely by voluntary participation in a variety of incentive programs. There are at least two unresolved issues. First, what is the intent of the Act? Are states expected to satisfy waterquality standards when available technologies are not sufficient to accomplish the task? Second, how do states control nonpoint sources to the MEP when those sources are largely exempt from regulation?

153 154 154 156 157 157 159 160 160

In this chapter, several nonpoint source management strategies are examined with respect to how ambiguities in the CWA either were addressed during implementation or have created as-yet insurmountable barriers to implementation. Particular attention is focused on policies for control of nonpoint sources of nutrients leading to hypoxia in receiving water and the management of urban stormwater. Cases include water-quality management for the Neuse River Basin, implementation of urban stormwater management in Jordan Lake watershed, North Carolina, control of nutrients in the Mississippi River Basin, and control of nutrients entering the Chesapeake Bay.

1.09.1 Nonpoint Sources and the CWA None of these issues were adequately addressed when fundamental changes to the federal water pollution control policy were made in 1972. That shortcoming is readily understood in the context of information about sources of pollution at the time. The Federal Water Quality Administration (FWQA) undertook the first national assessment of water quality in 1969. That assessment was largely a compilation of professional judgments by state water-quality officials. Results indicated that 33% of all stream mileage in the United States was polluted to some degree. One estimate in that assessment was that industrial sources accounted for 24% of degraded streams, municipal sources 22%, and agriculture 11% (USEPA, 1971). With that presumed factual basis, it is not surprising that Congress believed that strong regulatory limits on municipal and industrial sources would lead to substantial elimination of polluted waterways. Section 303(d)(1)(C) of Amendments to the Federal Water Pollution Control Act of 1972 (referred to as the CWA beginning in 1977) required establishment of total maximum daily loads (TMDLs) for all stream segments in which the application of technology-based effluent limits on municipal and industrial sources would be insufficient to implement the applicable water-quality standards. TMDLs had to account for seasonal variations in stream properties and a margin of safety to cover any lack of knowledge between effluent limitations and water quality. Section 302 stated that where technology-based standards were

153

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insufficient to satisfy water-quality standards, more stringent effluent limitations for point sources should be established at levels that can be reasonably expected to bring about compliance. USEPA introduced regulations to implement that portion of the act, using the following language (USEPA, 1975):

• • •

For each water quality segment, a total allocation for point sources of pollutants and a gross allotment for nonpoint sources of pollutants. A specific allowance for growth shall be included in the allocation for point sources and the gross allotment for nonpoint sources. The total of the allocation for point sources and the gross allotment for nonpoint sources shall not exceed the total maximum daily load.

The clear implication of the language in the statute and the regulations is that nonpoint sources could be addressed simply by reserving a portion of the TMDL for that purpose. The language did not address at least two possibilities: (1) that no economically reasonable available treatment technology for point sources would be sufficient to meet standards and (2) that loads from nonpoint sources on a stream segment would exceed the TMDL for that segment. Efforts to implement the policy quickly came up against realities of these two possibilities. The only option available to a state in this situation was to place a low priority on setting a TMDL for such a water body, thereby scheduling its development as far into the future as possible. When TMDL regulations were revised as of July 1987 (40 CFR 130.2), there was little change in the basic language. The loading capacity, defined as the maximum load of a pollutant on a stream segment that still satisfies related water-quality standards, is to be divided between a load allocation for nonpoint sources and background levels and a waste load allocation for point sources. These regulations went on to say that if best management practices (BMPs) to reduce nonpoint sources make more stringent load allocations for those sources practicable, then waste load allocations for point sources can be made less stringent. The regulations again failed to address the case where nonpoint source loads dominate the system and BMPs are insufficient to reduce actual loads below the TMDL. The latest version of TMDL regulations (40 CFR 130.7) states that TMDLs will be established ‘‘yat levels necessary to attain and maintain the applicable narrative and numerical water quality standardsy’’ taking into account all sources that are contributing to nonattainment of the standard. The directive to establish TMDLs does not carry with it any additional authority to control nonpoint sources. Amendments to the CWA in 1987, specifically the addition of Section 319, required states to take a number of additional steps to control nonpoint sources, including:

• •

preparation of an assessment report that identified navigable waters for which existing BMPs could not be expected to attain or maintain water-quality standards; identify BMPs and other measures to reduce to the MEP loads from each category of nonpoint sources; and

•

identify state and local programs for implementing BMPs and other measures.

States were then required to submit a management program ‘‘yto reduce pollutant loadingsy’’ from nonpoint sources ‘‘yto the maximum extent practicable.’’ Nowhere did the amendment state how much reduction was necessary. Section 319 did not require national technology standards or guidelines for nonpoint sources comparable to those for point sources in Section 302. Nor did it require any permitting, monitoring, inspection, and enforcement actions comparable to those for point sources as required under Section 402. State and USEPA officials were left to negotiate a mutually agreeable set of nonpoint source control measures, limited by whatever authorities and financial resources states and other federal agencies had to implement. Similar language was added in 1987 to Section 402(p) to address stormwater runoff from industrial activities and municipal separate storm sewer systems (MS4s). Although stormwater discharge was made subject to discharge permits, the operative provision for municipal permits is that dischargers are required to reduce pollutant loads to the MEP. That provision leaves open the question of what is the MEP. In particular, do the stormwater regulations apply to existing development as well as to new development? Portions of the stormwater provisions are rather straightforward, but, in addition to MET, other provisions are problematic. USEPA defined MS4s very broadly to include not only the conventional elements of urban stormwater management systems, but also roads with drainage systems, ditches, and man-made channels in urbanized areas that are owned by a state or any type of local government. At least two significant ambiguities have arisen from this definition and related requirements. First, urbanized areas do not necessarily coincide with boundaries of political jurisdictions to which the necessary regulatory authority has been delegated. Second, how would these provisions apply to large-scale real estate developments that convert rural areas to urbanized areas? In 1997 when EPA issued new policies for the TMDL process, it acknowledged that implementation of the program was moving at an unacceptably slow pace (USEPA, 1997). At the time, EPA regulations required each state to submit its list of impaired waters every 2 years and identify which of those would be scheduled for TMDL development over the following 2 years. The problem was that there was no schedule for all impaired segments. The revised policy urged states to develop schedules for all segments over a period of 8–13 years. Even then, implementation of nonpoint source controls was limited to the requirements in Section 319. When it issued proposed numerical nutrient criteria for the State of Florida in January 2010, EPA acknowledged that it could take many years before affected waters could be brought into compliance. Among many other provisions, the proposal asked for comments on Restoration Water Quality Standards that would allow achievement of water-quality standards in several phases over a period of up to 20 years so long as adequate progress was achieved at each stage (USEPA, 2010). This approach would allow considerable time to reach the final numerical criteria, but writing criteria does not grant additional authority to satisfy them.

Implementation of Ambiguous Water-Quality Policies

1.09.2 Intrastate Cases Two intrastate cases from North Carolina illustrate the use and limits of state authority to address problems of excessive nutrient loadings to lakes and estuaries, including urban stormwater runoff.

1.09.2.1 Neuse River Nutrient Management Strategy North Carolina’s Division of Environmental Management (DEM), the state’s water-quality management agency at the time, initiated a third-generation basinwide water-quality planning process in 1991 to coincide with 5-year renewals of the National Pollutant Discharge Elimination System (NPDES) discharge permits. The first of 17 basin plans was for the Neuse River Basin (NC Division of Environmental Management, 1993) shown in Figure 1. A supplemental classification of Nutrient Sensitive Waters had been assigned to the basin in 1983 to the Neuse River, and with that designation, all significant point sources had been required to meet a phosphorus effluent limit of 2.0 mg l 1 over the period of 1988–93. The 1993 report estimated that only 21% of the phosphorus and only 12% of nitrogen were coming from point sources. Agriculture accounted for approximately two-thirds of the balance. In compliance with USEPA policy, the report identified a wide range of state and federal programs to address nonpoint sources. It described 10 programs for agriculture, including the NC Agricultural Cost Share Program and provisions of the 1985 and 1990 Farm Bills. Four urban stormwater programs were discussed; runoff from construction was covered by the state sediment control act; and regulations on mining activities were discussed. Other programs for hydrologic modification, on-site disposal of wastewater, concentrated animal feedlots, solid waste, forestry, wetlands, and groundwater were also included in the mix.

The 1993 basinwide management plan, based on the CWA approach to nonpoint sources, failed to achieve its objective in late summer 1995 when massive fish kills occurred in the Neuse River estuary. Immediately thereafter, the North Carolina Environmental Management Commission (EMC) directed DEM to prepare a new plan to address deficiencies in the 1993 basin plan. The governor weighed in on the matter in the following March (Hunt, 1996). Draft administrative rules to implement the plan were sent to public hearing in May 1996 under the title of Neuse River Nutrient Sensitive Waters Strategy or the NSW Strategy. The state legislature provided legislative authority for a reduction of nitrogen by 30% from a 1991–95 baseline for both point and nonpoint sources (Session Law 1995 Chap. 572, ratified June 1966). This target was recommended by a special legislative committee that included knowledgeable academics and staff of DEM. Final rules were approved by the EMC in December 1997 (15A NCAC 2B.0232-0242 (Title 15, Chapter 2, Subchapter B, Sections of the North Carolina Administrative Code)), and specific state statutory authority was cited as the basis for these regulations. Several provisions of those rules established new initiatives to fill gaps in the basinwide management plan. For the first time, general agricultural operations were subject to waterquality regulations. Two options were made available to agricultural operators to satisfy the 30% reduction for nitrogen. They could act either individually to implement specified BMPs or collectively in a local (county) plan where some form of its trading among participants would be possible. A new program for certifying fertilizer applicators was established, covering all operations that applied fertilizer to 50 acres or more each year. Stormwater regulations were required for a list of municipalities and counties in the basin (they would later be covered by Phase II of USEPA regulations). Protection of existing riparian buffers was implemented in Section 233. A table of uses that are either exempt, allowable, allowable with mitigation, or prohibited applies to a

Durham

Wilson

Raleigh

Greenville

Goldsboro Kinston

Neuse River 0

Miles 10 20 30

Figure 1 Neuse River Basin, North Carolina.

40

155

New Bern

Neuse River Estuary

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50-ft-wide buffers directly adjacent to intermittent and perennial streams, lakes, ponds, and estuaries in the basin. A trading program for point sources was established in Section 02B.0235, building on experiences with the Tar RiverPamlico Sound trading program established in 1991. General powers delegated to the EMC by the state legislature were cited as the authority on which these administrative rules were adopted. EPA was in the process of formulating its policies for trading under the CWA. Unlike the Clean Air Act where national trading programs had been explicitly authorized to address acid rain, the CWA contained no mention of trading. EPA issued a policy statement on effluent trading in watersheds in January 1996, followed 4 months later by a draft framework for implementing the policy (USEPA, 1996). Basic principles for trading under the CWA were laid out in Chapter 2 of that document, but final rules were not formally proposed until proposed in the Federal Register until May 2002. Final trading rules were adopted in January 2003 (USEPA, 2003). Because many of the provisions in the NSW Strategy were new, they faced considerable obstacles and a few revisions before becoming effective. Considerable care was taken to involve stakeholders as the package of rules was being formulated, but public interest in finding a fix was very high. Not all stakeholders were enthusiastic about the rules adopted by the EMC. Under North Carolina’s Administrative Procedures Act, all administrative rules are subject to review by the Rules Review Commission (RRC), and no rule can become effective until after the next session of the legislature except in special cases where temporary rules are permitted. With very strong political support to get the rules in place, the legislature allowed all but the initial riparian buffer rule to become effective. This rule was slightly modified to meet objections by one group of stakeholders.

While the rules in Section 235 and 238 established a regulatory program on agriculture for the first time, they were limited to the installation of BMP and an annual reporting of what BMPs were in place and how much acreage was subject to BMPs. There is very limited instream monitoring to determine effectiveness of the rules, and inspections of operations subject to the rule are limited to sparse visits by agricultural agencies.

1.09.2.2 Jordan Lake Stormwater Rules A similar strategy was developed for the protection of B. Everett Jordan Lake, a critical source of public water supply in the Research Triangle of North Carolina. An important difference between the Neuse River and Jordan Lake nutrient loads is the relative importance of urban stormwater from existing development, a difference that was addressed in formulating a management plan for the lake. In the wake of actions in 1996 to address nutrient problems in the Neuse River estuary, the state legislature passed the Clean Water Responsibility Act (CWRA; also referred to as House Bill 515) in August 1997 to protect inland lakes that are designed as nutrient sensitive by the EMC. CWRA established default limits in the form of maximum concentration values for phosphorus and nitrogen in effluents from dischargers above such lakes. It also provided that alternate mass load limits could replace default values if they were based on a calibrated nutrient response model. Impacts estimated by mass loading would have to show compliance with the waterquality standard of 40 mg l 1 of chlorophyll-a. Primary targets of the CWRA were Jordan Lake in the Cape Fear Basin, shown in Figure 2, and Falls Lake in the Neuse Basin, both US Army Corps of Engineers multipurpose

Kernersville

Durham Burlington Greensboro Chapel Hill Jordan

Miles 0

5

10

15

20

Figure 2 Jordan Lake Drainage Area, Cape Fear Basin, North Carolina.

Lake

Implementation of Ambiguous Water-Quality Policies

reservoirs with a combined safe yield of about 165 million gallons per day for public water supplies in the area. Before Jordan Lake was found to be in violation of the water-quality standard in 2002, local governments in the watershed had already initiated development of a nutrient response model. A series of nutrient delivery models for point sources and nonpoint sources were developed over the period of 2001–03. A model of lake quality was delayed by errors in data collection, and a final report was not delivered until February 2005. Twenty-one stakeholder meetings were held in the process of developing the model (North Carolina Division of Water Quality, 2006). An initial draft of an NSW and TMDL strategy was submitted by the NC Division of Water Quality (DWQ) to the EMC in April of 2005, and draft rules were adopted by the Commission in November 2005. A set of technical issues were hashed out in meetings with stakeholders in 2006, before the proposed regulations were sent to formal public hearings in March 2007. The EMC approved a revised set of rules in May 2008, and rules that were passed were approved by the RRC in November 2008 in time to be considered by the 2009 session of the state legislature. Jordan Lake rules are very similar in structure to the Neuse Rive NSW Strategy with one very important exception. All prior regulations in NC directed toward control of pollutants in urban stormwater had been limited to the management of runoff from new development. However, in the process of making load allocations for Jordan Lake, DWQ could not find a feasible and equitable distribution of loads among sources without requiring reductions from existing urban development. An existing development rule was included as part of the management program. Several major urban areas in the watershed, including Burlington, Chapel Hill, Durham, and Greensboro would be affected by the proposed rule. The EMC proposed an adaptive approach to reducing loads from existing development. Municipalities and counties were given

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3 years and 6 months from the effective date of the rule to conduct feasibility studies and prepare implementation plans. Following approval by DWQ, they would have another 4 years and 6 months to initiate implementation (North Carolina Division of Water Quality, 2008). Affected municipalities and counties sought relief from the existing development rule in the 2009 session of the legislature. The legislature responded by disapproving the EMC’s existing development rule, but included much of the content of the rule in the same statute (Session Law 2009-216). The primary change was the schedule for implementation. That action concluded a 10-year process for revising protection strategies adopted in the Cape Fear Basinwide Water Quality Plan approved by the NC EMC in October 1996.

1.09.3 Interstate Nonpoint Management In the Neuse and Cape Fear River Basins, both general and explicit state authorities over nonpoint sources were used to fill gaps in the CWA in efforts to bring waters in those basins into compliance with water-quality standards. Interstate problems are much more intractable, especially in the absence of a strong federal role.

1.09.3.1 Mississippi River Basin and Hypoxia in the Gulf of Mexico An initiative to address the growing problem of hypoxia in the Northern Gulf of Mexico by reducing nutrient inputs from the Mississippi River Basin is a case in point. The areal extent of the hypoxic zone, located along the Louisiana coast as shown in Figure 3, is highly variable from 1 year to the next, but the trend is upward, and over the period 1996–2000 the average was over 14 000 square kilometers (about 5500 square miles).

Upper Mississippi

Missouri

Ohio

Arkansas Red Tennessee Lower Mississippi Atchafalaya

Hypoxic Zone Gulf of Mexico

Figure 3 Mississippi–Atchafalaya River Basin and Hypoxic Zone in the Gulf of Mexico.

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Concerns about expansion of the zone led to the formation of the Mississippi River/Gulf of Mexico Watershed Nutrient Task Force (MR/GMWNTF) in 1997 comprised of a federal interagency working group and representatives of states and tribes. In November 1998, Congress passed the Harmful Algal Bloom and Hypoxia Research and Control Act of 1998 (HABHRCA, Title VI of P.L. 105-383), gave statutory authority to the task force, and directed it to conduct a scientific assessment of the problem and develop an action plan to control it. The Task Force was administered by the Committee on Environment and Natural Resources of the National Science and Technology Council. The first Action Plan produced in 2001 established the following goal (MR/GMWNTF, 2001): By the year 2015, subject to the availability of additional resources, reduce the 5-year running average areal extent of the Gulf of Mexico hypoxic zone to less than 5000 square kilometers through implementation of specific, practical, and cost-effective voluntary actions by all states, tribes, and all categories of sources and removals within the Mississippi/Atchafalaya River Basin to reduce the annual discharge of nitrogen into the Gulf.

An extensive body of scientific material was produced by the Task Force. Estimates of nutrient and sediment loads, their sources, and amounts delivered to the Gulf have been developed by a group of researchers at the United States Geological Survey (Alexander et al., 2008). Nitrogen and phosphorus delivered to the Gulf arrive from numerous tributary basins as shown in Table 1. Table 1 Percentage distribution of nutrients delivered to the Gulf of Mexico by Source Basin Basin

Upper Mississippi Missouri Central Mississippi Ohio and Tennessee White Arkansas Red Lower Mississippi and Atchafalaya Total

• •

•

Phosphorus

6.5 12.8 25.0 36.1 1.6 4.5 1.0 12.6

5.4 9.3 21.3 35.8 2.6 5.0 1.9 18.7

100.0

100.0

establishment of several subbasin committees; development of an integrated monitoring, modeling, and research strategy; additional monitoring of the hypoxic zone; increased assistance to agricultural producers through US Department of Agriculture programs to restore wetlands, add stream buffers, and install other best management practices; and completion of a major reassessment of the science that supports action items.

Other action items were not implemented, however, substantially impeding reductions in nutrient loads. Among the impediments cited in the report are:

• • •

failure to develop an integrated federal budget to support voluntary nutrient reduction; not only a failure to expand a long-term monitoring program, but also discontinuance of some stations; and the slow pace of development of sub-basin strategies.

Lack of progress on reducing nutrients may be attributable to several factors. First, although there may be strong desires among Gulf states and at the federal level to reduce hypoxia in the Gulf, upstream states lack a compelling interest to impose regulations on nonpoint source generators. Second, competing priorities in the Gulf states, particularly recovery from Hurricanes Katrina and Rita, could have delayed funding of the program. Third, the CWA is ambiguous on authority of

Percentage distribution of nutrients delivered to the Gulf of Mexico by type of source

Source

Nitrogen Mississippi

Urban and population-related sources Atmospheric deposition Crops Pasture/rangeland Forest Shrub and barren lands Total

• •

Percent of total Nitrogen

Table 2

Activities from which loads delivered to the Gulf are generated are shown in Table 2. Point sources account for less than 10% of the nitrogen load; they account for only 10–12% of phosphorus loads. Crops and pasture/rangelands account for approximately 70% of nitrogen and about 80% of phosphorus. The system is clearly dominated by nonpoint sources. Numerous interstate cooperative efforts among both governmental and nongovernmental agencies and organizations have been organized to address a range of water-resource issues along the Mississippi River, including water quality, navigation, and public water supply. Some of the more significant organizations are covered by a report of the National Research Council (NRC, 2008). MR/GMWNTF, with its 2001 Action Plan and its 2008 update (MR/GMWNTF, 2008) developed under authority of HABHRCA of 1998 and 2004, is the most effective action taken to date. The 2008 Action Plan reported progress on action items in the 2001 Plan, including:

Phosphorus Atchafalaya

Mississippi

Atchafalaya

9.1 16.2 65.6 5.0 4.1 0.2

8.9 18.0 62.2 5.6 5.2 0.2

12.3

10.6

43.1 37 7.5 0.4

40.0 39.4 9.2 0.97

100.0

100.0

100.0

100.0

Implementation of Ambiguous Water-Quality Policies

USEPA to take or even threaten to take regulatory action to achieve the goal. HABHRCA did not grant any additional authority to implement the 2001 Action Plan, and, as cited above, the Task Force conditioned the goal of reducing the hypoxic zone on the availability of additional resources and voluntary actions by all states. Several provisions of the CWA explicitly address interstate cooperation on nonpoint sources. These include:

• • •

Section 103(a) – Interstate Cooperation and Uniform Laws Section 319 – Nonpoint Source Programs Section 320 – National Estuarine Program.

Section 103(a) is a very brief admonition to USEPA to ‘‘yencourage cooperative activities by the Statesy’’ to manage pollution, promote uniform state laws, and encourage states to form interstate compacts when necessary to address cross-border effects. Specific authority is granted in Section 301(b) for two or more states to form compacts and agreements to control pollution. Section 319(g), part of the 1987 amendments, gives a state that does not meet water-quality standards due to nonpoint sources in other states the right to petition USEPA to convene a management conference. If USEPA finds that available information is sufficient to support the petition, tributary states are to be notified and a management conference is to be convened. USEPA is also given authority to initiate a management conference without a petition. The purpose of a conference is to develop an agreement among the participating states to improve water quality, but it is quite unclear as to what happens if the states do not reach agreement. The language is: ‘‘To the extent that the States reach agreement through such conference, the management programs of the Statesywill be revised to reflect such agreement.’’ Section 319(g)is silent on what happens in the absence of an agreement among the states. The National Estuary Program, authorized under Section 320 of the CWA, similarly directs USEPA to convene management conferences for the protection of designated national

estuaries in which water-quality standards are not being met. Estuarine management conferences are intended to assess water-quality trends, identify causes of pollution, establish relationships between pollution loads and water quality, develop comprehensive management plans, and identify federal financial assistance. The Barataria–Terrebonne Estuarine Complex along the Louisiana coast is the primary system that would be covered by this program. Section 303 of CWA – Water-Quality Standards and Implementation Plans does not mention other states in its requirements for listing of impaired waters and setting of TMDLs. Nonetheless, the Committee on the Mississippi River and the CWA (NRC, 2008) argued that USEPA has interpreted the CWA as imposing obligations on each state to protect downstream water quality in other states when setting TMDLs. A fundamental barrier to implementation of these authorities is that neither CWA nor HABHRCA grants USEPA or other federal agency any authority to set enforceable limits on nonpoint sources at levels sufficient to attain and maintain water-quality standards. In all cases, implementation depends on the will of tributary states to adopt effective nonpoint source programs, and USEPA is obligated to accept a state management plan if the state satisfies the criterion to identify BMPs and other measures to reduce to the pollution loads to the MEP.

1.09.3.2 Chesapeake Bay Program Urban growth, more intensive agricultural operations, and other factors in the watershed resulted in significant deterioration of water quality and related ecosystems in the Chesapeake Bay. Among other undesirable outcomes was widespread hypoxia due to excess nutrient loads. As shown in Figure 4, areas that drain to the Chesapeake Bay cover portions of several states. In 1983, 1987, and again in 2000, several states in the Chesapeake Bay watershed, the District of Columbia, the Chesapeake Bay Commission, and the US EPA entered into agreements to form the Chesapeake Bay

New York

Pennsylvania Basin boundary NJ

MD

West Virginia

Figure 4 Chesapeake Bay Drainage Area.

159

DE Virginia

160

Implementation of Ambiguous Water-Quality Policies

Table 3

Distribution of nutrient loads delivered to Chesapeake Bay by source

Source of nitrogen

Percent of total load

Source of phosphorus

Percent of total load

Septic Municipal and industrial wastewater Manure (agriculture) Chemical fertilizer (agriculture) Chemical fertilizer (nonagriculture) Atmospheric

4.3 18.9 17.8 15.4 10.1 33.5

Manure (agriculture) Chemical fertilizer (agriculture) Municipal and industrial wastewater Natural (wildlife and forest) Other fertilizer

27 18 22 3 30

Source: Chesapeake Bay Program, www.chesapeakebay.net/tribtools.htm#allocations.

Table 4 Direct funding provided by the federal agencies, states, and District of Columbia, fiscal years 1995 through 2004, in millions of constant 2004 dollars Federal agencies Department EPA Department Department Department

of Defense of Agriculture of the Interior of Commerce

Total – federal agencies

States 355.4 253.7 230.4 77.4 55.5

Maryland Virginia District of Columbia Pennsylvania

1862.4 752.6 41.8 28.1

972.4

Total – all states

2684.8

Source: United States Government Accountability Office, 2005.

Program (CBP) with the intent of restoring the Bay’s water quality and ecosystem. Maryland, Pennsylvania, and Virginia account for 88% of the nitrogen load and 87% of phosphorus. Distributions of sources of nitrogen and phosphorus that are delivered to the Bay are given in Table 3. As part of the 1987 agreement, the signatories committed to reduce 40% of nutrient loads that were controllable. Tributary-specific nutrient reduction strategies were adopted in 1992. The 2000 agreement asserted that where the 1992 goals had not been achieved, additional steps would be taken (CBP, 2000). Direct federal and state expenditures over the decade 1995–2004 are shown in Table 4. Federal agencies spent an average of nearly $100 million (in constant 2004 dollars) a year, and the three states and District of Columbia spent about $270 million a year (United States Government Accountability Office, 2005). Current estimates of distributions of nitrogen and phosphorus loads entering the Bay indicate that runoff from agriculture and urban areas account for large percentages of total loads. Nearly 60% of nitrogen and 80% of phosphorus loads are attributed to those sources (CBP, 2009). More recently, President Obama, citing lack of adequate progress toward restoration, issued Executive Order (EO) 13508 in May 2009, directing federal agencies to take a more active program to protect and restore water-quality and related ecosystems. Among the key challenges listed in the EQ are efforts to: (1) strengthen permit conditions for CAFOs and urban stormwater runoff and (2) enhance federal and state initiatives for conservation practices on agricultural operations (Federal Leadership Committee, 2009). USEPA also issued a notice of intent in September 2009 to establish a Bay-wide

TMDL for nutrients and sediments (USEPA, 2009). Preliminary targets for nitrogen and phosphorus loads have been established pursuant to the EO (Early, 2009). Maryland, Pennsylvania, and Virginia were allocated 88% of the nitrogen target and 89% of the phosphorus target. Even with these initiatives coming on authority of the President, it is far from clear as to how much can be achieved given ambiguities in the CWA. CAFO reductions may be more predictable because of existing permit authority, but controls on urban runoff will still be limited by the MEP criterion. All of the agricultural runoff programs cited in the EO are voluntary. Regulatory authority over urban stormwater and nonCAFO agriculture will be based on authorities of the states and the District of Columbia. In this case, the three states that are big contributors have a compelling interest and a history of strong political support for those policies, but it remains questionable as to how far that authority extends and how much political support there will be as reductions from these sources approach the limits of MEP.

1.09.4 Summary and Conclusions The several cases described in this chapter point to a conclusion that, in the absence of a compelling state interest to improving water quality, ambiguities in the CWA are likely to present strong impediments for achieving the goals of the Act. The goal of the Act is protect and enhance water quality sufficient to support designated uses of all stream segments, the lowest acceptable use being propagation of fish and aquatic ecosystems. Few ambiguities exist when it comes to managing point sources; effluent limits must be set to satisfy waterquality standards, and state boundaries do not present an insurmountable barrier. TMDLs set by one state must consider effects on waters in downstream states. However, CWA nonpoint sources require only that states employ management practices that reduce loads to the maximum extent practical, a judgment call that is left to the states. A similar requirement applies to stormwater runoff from urbanized areas. In the Neuse River, Jordan Lake, and Chesapeake cases discussed in this chapter, compelling local and regional interests have led to invocation of state authority to shore up gaps in federal legislation. It may be argued that Congress intended just that, but, in these cases, states have followed very closely the language of the CWA. States have made judgments as to what BMPs satisfy the criterion of MEP without effective

Implementation of Ambiguous Water-Quality Policies

monitoring, inspection, and enforcement procedures to insure that water-quality standards will be attained and maintained. While it may not be realistic to expect compliance with waterquality standards over short time horizons given existing land uses and agricultural and urbanization practices, ambiguities in present federal and state policies about nonpoint sources leave the public poorly informed as to what is possible and the extent to which management programs are being effective.

References Alexander RB, Smith RA, Schwarz GE, Boyer EW, Nolan JV, and Brakebill JW (2008) Differences in phosphorus and nitrogen delivery to The Gulf of Mexico from the Mississippi River Basin. Environmental Science and Technology 42(3): 822--830. CBP (Chesapeake Bay Program) (2000) ‘‘Chesapeake 2000’’, the 2000 Partnership Agreement. http://www.chesapeakebay.net/content/publications/cbp_12081.PDF (accessed April 2010). CBP (Chesapeake Bay Program) (2009) Bay Barometer: A Health and Restoration Assessment of the Chesapeake Bay and Watershed in 2008. http:// www.chesapeakebay.net/news_baybarometer08.aspx?menuitem=34917 (accessed April 2010). Early WC (2009) Acting Regional Administrator, Region III, Letter to L. Preston Bryant, Jr., Virginia Secretary of Natural Resources, 3 November 2009. http://www.epa.gov/ region3/chesapeake/bay_letter_1209.pdf (accessed April 2010). Federal Leadership Committee (2009) ‘‘Executive Summary for Draft Reports Addressing Key Challenges to Chesapeake Bay Protection and Restoration’’. Reports prepared in response to Section 202 of Executive Order 13508, 9 September 2009. Hunt JB (1996) Letter to David H. Moreau, Chairman, North Carolina Environmental Management Commission, 12 March. MR/GMWNTF (Mississippi River/Gulf of Mexico Watershed Nutrient Task Force) (2001) Action Plan for Reducing, Mitigating, and Controlling Hypoxia in the Northern Gulf of Mexico, Washington, DC. http://www.epa.gov/msbasin/pdf/ actionplan2001.pdf (accessed April 2010). MR/GMWNTF (Mississippi River/Gulf of Mexico Watershed Nutrient Task Force) (2008) Gulf Hypoxia Action Plan 2008. Washington, DC: Office of Wetlands, Oceans, and Watersheds, United States Environmental Protection Agency.

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NC Division of Environmental Management (1993) Neuse River Basinwide Water Quality Management Plan, Raleigh, NC. North Carolina Division of Water Quality (2006) Jordan water supply nutrient strategy and rules. Report to the NC Environmental Management Commission, 11 January 2006. North Carolina Division of Water Quality (2008) ‘‘Jordan Lake nutrient strategy’’. http:// portal.ncdenr.org/web/wq/ps/nps/jordanlake (accessed April 2010). NRC (National Research Council) (2008) Mississippi River Water Quality and the Clean Water Act. Washington, DC: National Academy Press. United States Government Accountability Office (2005) Chesapeake Bay program: Improved strategies are needed to better assess, report, and manage restoration progress. Report to Congress. Washington, DC: United States Government Accountability Office. USEPA (United States Environmental Protection Agency) (1971) The Cost of Clean Water, Volume I, Annual Report (4th) to Congress in Compliance with Section 26(a) of the Federal Water Pollution Control Act, Senate Document 92-23. Washington, DC: United States Government Printing Office. USEPA (United States Environmental Protection Agency) (1975) Preparation of water management plans. Federal Register 40(230): 55345--55346 (28 November 1975) USEPA (United States Environmental Protection Agency) (1996) Draft Framework for Watershed Based Trading. EPA Report No. 800 R 96 001. Washington, DC: USEPA (30 May 1996). USEPA (United States Environmental Protection Agency) (1997) ‘‘New Policies for Establishing and Implementing Total Maximum Daily Loads (TMDLs),’’ Memorandum from Robert Perciasepe, Assistant Administrator to Regional Administrators Regional Water Division Directors, 8 August 1997. http:// www.epa.gov/OWOW/tmdl/ratepace.html (accessed April 2010). USEPA (United States Environmental Protection Agency) (2003) Water Quality Trading Policy. Washington, DC: Office of Water. (13 January). USEPA (United States Environmental Protection Agency) (2009) Clean Water Act Section 303(d): Preliminary Notice of Total Maximum Daily Load (TMDL) Development for the Chesapeake Bay. Federal Register 74(179): 47792--47794 (17 September 2009). USEPA (United States Environmental Protection Agency) (2010) Water Quality Standards for the State of Florida’s Lakes and Flowing Waters: Proposed Rule. Federal Register 75(16): 4291 (26 January 2010).

1.10 Predicting Future Demands for Water B Dziegielewski and DD Baumann, Southern Illinois University Carbondale, Carbondale, IL, USA & 2011 Elsevier B.V. All rights reserved.

1.10.1 1.10.1.1 1.10.1.2 1.10.2 1.10.2.1 1.10.2.2 1.10.2.2.1 1.10.2.2.2 1.10.2.2.3 1.10.2.2.4 1.10.3 1.10.3.1 1.10.3.1.1 1.10.3.1.2 1.10.3.2 1.10.3.2.1 1.10.3.2.2 1.10.3.2.3 1.10.3.3 1.10.3.3.1 1.10.3.3.2 1.10.4 1.10.4.1 1.10.4.2 1.10.4.2.1 1.10.4.2.2 1.10.4.2.3 1.10.4.3 1.10.4.3.1 1.10.4.3.2 1.10.4.4 1.10.4.4.1 1.10.4.4.2 1.10.4.5 1.10.4.5.1 1.10.4.5.2 1.10.5 1.10.5.1 1.10.5.2 1.10.5.3 1.10.5.4 1.10.6 References

Water Supply and Demand Changing Objectives of Water-Supply Development Emergence of Water Conservation Water-Use Data Definitions and Measurement of Water Use Accessibility of Data on Water Use Public-supply sector data Industrial and commercial sector data Power generation sector data Irrigation sector data Water-Demand Relationships Theoretical Models of Water Demand Derived demand of producers Final demand of consumers Empirical Models of Water Use Configuration of data sets Functional forms and model parameters Elasticities with respect to major determinants of water use Other Water-Use Relationships Cooling water requirements Supplemental irrigation water requirements Demand Forecasting Techniques Forecasting Principles and Criteria Forecasting Models and Procedures Time trend forecasting Water requirement forecasts Demand forecasts Dealing with Forecast Uncertainty Model-dependent prediction intervals Dealing with forecast assumptions error Forecasts with Conservation End-use accounting system Baseline and restricted forecasts Forecasting Software: The IWR-MAIN Program Model structure and procedures IWR-MAIN conservation forecasts Example of a Regional Multisector Forecast Water-Use Relationships Effects of Key Forecast Assumptions Effects of Future Climate Forecast Summary Conclusion

1.10.1 Water Supply and Demand Water is an essential natural resource, which plays a vital role as input into many economic activities adding to the quality of human life and supporting the health of ecological systems. Water is also a commodity which has an economic value in all its competing uses as has been recognized by the main international conferences on water (e.g., ICWE, 1992;

163 163 163 164 164 165 165 165 166 166 166 166 167 168 169 169 170 171 172 172 173 175 175 175 175 176 177 177 178 178 178 179 179 179 179 180 181 182 182 183 185 186 187

UNCED, 1992). Despite this, the importance of water supply may not be widely appreciated by the general public because only some water resources and water uses are easily noticed while others are not. Surface water resources such as rivers and lakes are highly visible and well recognized for their cultural and amenity values as well as for their important functions in outdoor recreation and transportation. Less recognized by the general public is the portion of water

163

164

Predicting Future Demands for Water

resources – groundwater – which in the United States (the lower 48 states) actually represents three-fourths of all freshwater storage (Dziegielewski and Kiefer, 2006). Similarly, some human uses of water are easily noticed while others are not. Use of water for hydropower production or for irrigation can be easily seen and appreciated, while large water flows to urban and industrial users are usually hidden in underground pipes. For example, the network of public water supply pipes in the US carries an average flow of 67 000 cubic feet per second (cfs) – this is equivalent to the average discharge of a large river. The flow of water for thermoelectric cooling is even greater; in the US in the year 2000, it reached 302 000 cfs – more than the average annual discharge of the Ohio River. Given the overall importance of water, it is understandable that the long-term adequacy of water supply is a major national concern in many countries. However, in order to assess the future adequacy of supplies, it is necessary to determine the amount of water that is used currently and the amount that will be demanded in the future. These demands have to be compared with the future availability of water in existing and potential sources of supply. To do this, water-supply planners need appropriate tools for quantifying future water demands and assessing the effects of future climate and other factors on both water demand and water availability. Credible long-term forecasts of water demand are essential to all type of planning. Without such forecasts water planners cannot achieve an efficient allocation of water supplies among competing uses or ensure their long-term sustainability.

1.10.1.1 Changing Objectives of Water-Supply Development The traditional approach to water-supply planning has evolved during the past century, as cities, industries, and irrigation districts have expanded their water-supply infrastructure (Blake, 1956; Grigg, 1986). In urban areas, municipal water-supply agencies considered their responsibility to be on-demand delivery of sufficient quantities of drinking quality water while maintaining adequate pressure for consumption and fire protection. Even very high expenditures on the development of waterworks could be justified because adequate community water supply was considered an essential service that ensured public health and safety, economic activity, and a general community well-being. Also, accurate long-term forecasts of water demand were not very important because rapid urban growth made it possible to add large increments of water-supply capacity without the risk that it would remain unused for long periods of time. The longstanding practice of augmenting water supply by securing reliable sources, protecting water quality at the source, and developing adequate facilities for transmission, treatment, and distribution was based on several advantages of this strategy. These were summarized by Platt (1993). Specifically, somewhat distant hinterland sources provided a cheap and abundant supply of high-quality water without the need for treatment. Cities in Northeast and Northwest of the US relied on such sources prior to the adoption of water chlorination. In addition, storage reservoirs at relatively high elevations permitted gravity-flow transmission of water without the need for pumping. Also, urban governments could acquire

water rights and watershed lands for source protection at a minimum cost, often under condemnation powers, while the hinterland regions offered little political resistance to water exports. Finally, large professionally managed sources of supply offered significant economies of scale. The advantages of the hinterland sources provided little incentive for a careful matching of water supplies with water demands.

1.10.1.2 Emergence of Water Conservation Since the 1970s, the viability of the traditional supply augmentation approaches has begun to decline gradually because of several obstacles (Platt, 1993). The hinterland regions have begun to offer some political and legal resistance to water exports. Environmental legislation and interests have introduced significant barriers to a continuing expansion of offstream uses of water for urban and agricultural purposes. Major droughts over the last 40 years have also contributed to the increased competition for water supplies between urban and agricultural interests. Other factors such as the depletion and contamination of groundwater sources, difficulties in financing major construction programs, especially those sponsored by the federal government, and increasing costs of water treatment for regulated contaminants have made the traditional options of supply augmentation less viable. As a result, the range of options has expanded to include both some unconventional supply-side alternatives and the opportunities for modifying the growth in water demand. The introduction of demand management alternatives represents an important change in water-supply planning. In the early 1980s, the growing attractiveness of long-term demand management measures began to catch the attention of urban water-supply agencies (Boland et al., 1982; Dziegielewski and Baumann, 1992). Demand reduction programs allowed some agencies to balance future supply and demand at a cost that is below the economic, social, and environmental cost of new supply development and thus result in net benefits to society (Dziegielewski et al., 1983). Baumann et al. (1984) developed a practical definition of long-term water conservation as ‘‘yany beneficial reduction in water use or in water losses.’’ The authors pointed out that a water management practice constitutes conservation if it conserves a given supply of water through reduction in water use (or losses) and if it results in a net increase in social welfare where the resources used have a lesser value than those saved (p. 431). In other words, the beneficial effects of the reduction in water use (or loss) must be considered greater than the adverse effects associated with the commitment of other resources to the conservation effort. This definition provided an important guidance for long-term conservation; however, it could not be easily applied to short-term conservation measures which are usually aimed at curtailing water demand during a drought. Temporary restrictions on water use are usually undertaken in order to prevent adverse impacts of severe shortages in the future if the drought continues and their outcomes cannot be easily analyzed through benefit–cost analysis. Other marked enhancements in theory and practical knowhow for planning and evaluation of demand management alternatives have also been achieved (Dziegielewski et al., 1993). These were followed by the development of computer

Predicting Future Demands for Water

software programs for disaggregate water-use forecasting, for the analysis of demand reduction alternatives, for optimization of long-term water management plans, and for the monitoring of water demands over time (Dziegielewski and Boland, 1989; Dziegielewski, 1993; Baumann et al., 1998). This chapter provides a review of key water-demand concepts and presents analytical methods for quantifying demands by different user groups, forecasting future demands and analyzing demand management alternatives. The chapter begins with the discussion of the measurement and analysis of water use and related concepts.

1.10.2 Water-Use Data Our knowledge of water demands necessarily derives from the measurement and estimation of water use. In practice, it is impossible to know precisely all water uses – there are many different types of water users and specific purposes of use and only some uses are metered. Instead, various estimation methods are usually employed to determine the quantity of water use. Because water use depends on many factors, the analysis of water demands requires data on those factors. This chapter describes the structure of water demand and identifies the factors affecting water use as well as methods for analyzing the available water-use data.

1.10.2.1 Definitions and Measurement of Water Use From the hydrologic perspective, water use is a part of the water budget. At the most general level, water use can be defined as all water flows that are a result of human intervention within the hydrologic cycle. Accordingly, all water uses can be divided into in-stream and off-stream uses. In-stream use represents water that is used, but not withdrawn, from a natural water source for purposes such as hydroelectric power generation, navigation, water-quality improvement, fish propagation, and recreation. Off-stream use represents water withdrawn or diverted from a groundwater or surface water source for public water supply, industry, irrigation, livestock, thermoelectric power generation, and other uses (Hutson et al., 2004). The term ‘water withdrawal’ is used to designate the amount of water that is taken out from natural water sources such as lakes, rivers, or groundwater aquifers. The difference between the amount of water withdrawn and water returned to the source (also referred to as discharge) is usually taken to represent consumptive use. This is the ‘‘part of water withdrawn that is evaporated, transpired, incorporated into products or crops, consumed by humans or livestock, or otherwise removed from the immediate water environment’’ (Hutson et al., 2004). The part of amount withdrawn and returned back to the source is called nonconsumptive use. The quantity of water consumed is utilized in calculating regional annual and monthly water budgets, and represents a measure of the volume of water that is not available for repeated use. While a major portion of water withdrawn for purposes such as public water supply, power generation, and industrial use represents nonconsumptive use, these withdrawals can have significant impacts on water resources and other uses of

165

water. For example, water withdrawn from an aquifer and then returned into a surface water body may have a positive impact on streamflow or lake water levels, but a negative impact on the source of groundwater. Similarly, water withdrawn from a river for public water supply must be continuously available at the intake but not for withdrawal for other uses upstream or immediately downstream from the intake. A more restrictive definition of water use refers to water that is actually used at a specific site or for a specific purpose. Individual residential or commercial buildings, industrial facilities, and other locations can obtain water from their own sources of supply or through connections to a public or private distribution system. Individual users of water within a defined geographical area can be classified into different categories and their combined use can be summed up into broader categories, or user sectors. For example, the United States Geological Survey compiles data on water withdrawals for individual counties for eight categories of users: public supply, domestic, industrial, commercial, mining, power generation, livestock, and irrigation. Some of the categories are further subdivided based on the purpose of use. For example, irrigation use is subdivided into cropland and golf course irrigation. Similarly, public-supply use can be subdivided into residential, commercial, industrial, and public categories with each category further subdivided into two or more subcategories (e.g., residential single-family and residential multifamily). Measurement of water use can take place at the point of withdrawals or at the point of water use. Also, some measurements could be taken at the point of water treatment or along water transmission routes. Direct measurements of water volume being transmitted over a given period of time are made by meters which register the volume of flow (such as displacement meters) or by measuring and recording instantaneous flow (such as in Venturi meters). The measurements of water use are reported as water volume per unit of time. The volumetric units include cubic meters, cubic feet, gallons and liters, and their decimal multiples. In some cases, composite volumetric units (e.g., acre feet) or units of water depth (e.g., inches or centimeters of rainfall) may be used. The time periods used may include a second, minute, hour, day, month, and year. Because the annual and monthly volumes of water use generally involve large numbers, the numerical data on water use are often reported as the average daily quantities used. Two popular units are thousand cubic meters per day (1000 m3 d1) and million gallons per day (mgd). Also, in order to make the estimates of water use easy to comprehend and to make meaningful comparisons of water use for various purposes (and various users), the annual or daily quantities can be divided by some measures of size for each purpose of use. The result is an average rate of water use such as gallons per capita per day (gpcd), gallons per employee per day (ged), or other unit-use coefficients. Finally, it is important to note that the reported quantities of water use can be in the form of direct measurements obtained from water meters or they may be estimates. Estimates of water use that are derived from the measurements of water levels in storages or from pumping logs are generally more accurate than those derived from related data on the volume

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Predicting Future Demands for Water

of water-using activity. For example, the estimates of water use for hydroelectric power generation may be obtained by multiplying the amount of generated power by a water-use coefficient. In analyzing water demands, it is important to recognize the sources and nature of the data on water use. For example, when deriving statistical water-demand relationships, it makes little sense to use data which are estimates based on the volume of water-using activity or other correlates of water use. Only actual measurements of water volumes, which are withdrawn or used over time, can accurately capture the temporal and spatial variability of water demand and provide a basis for deriving econometric or other models of water use.

1.10.2.2 Accessibility of Data on Water Use The availability of data on water use depends on user sector. The best data are available for public-supply sector (also referred to as municipal and industrial use). Data for other sectors, which typically rely on self-supplied sources of water, are usually less precise and their accuracy depends on whether the withdrawals of water are regulated and are required to be metered or if only an annual reporting of the estimated quantities is required. The discussion of the available data for the different sectors is given in the following.

1.10.2.2.1 Public-supply sector data Water use in the public-supply sector can be characterized with respect to: (1) the demands of different types of customer classes (e.g., single-family residences, hotels, food-processing plants, etc.); (2) the purposes for which water is used (e.g., end-uses such as sanitary needs, lawn watering, or cooling); and (3) the seasonal variations in water use. The breakdown of total urban water use into customer groups, specific end-uses, and seasons can serve as a basis for modeling water demands and for conducting impact evaluations of water conservation programs. Generally, there are three types of water measurement records that are maintained by public water-supply utilities: (1) water production records (i.e., amounts of water pumped into the distribution system); (2) water billing records (i.e., records which detail each customer’s account activity); and (3) water sales records (i.e., summaries of total water sales or water sales by customer groups). Water production records show the amount of water pumped from treatment plants and are typically generated daily or hourly. Production data can be used for analyzing: (1) unaccounted water use (comparing production with water sales data); (2) impacts of water-use restrictions on total water demand; (3) relationships between total water demand and weather conditions; and (4) peak water use during different time intervals (e.g., peak day, peak hour, day of week, etc.). Table 1 (Dziegielewski and Chowdhury, 2008) shows an example of water production records for a sample of cities and water-supply systems in the Greater Chicago Area in Illinois. The last column shows the calculated rate of usage in gpcd. Water billing records represent the individual customer account data which are generally maintained by retail watersupply agencies. The individual record usually includes: (1) name and address of account holder, (2) type of account

Table 1 Examples of water production and purchase data by public water supply systems in selected communities in Northeastern Illinois in 2005 Community or system name

Aurora Bedford Park Belvidere Central Lake Co. JAWA Chicago Crystal Lake DeKalb DuPage Water Com. Elgin Evanston Glencoe Hammond WSS Highland Park Joliet Kankakee Aqua Illinois Lake County PWD Lake Forest Morris North Chicago Northbrook Northwest Sub. M. JAWA Oak Lawn Oswego Waukegan Wilmette Winnetka Total study area

Reported production in mgd

Estimated population served

Per capita production in gpcd

18.1 25.4 3.7 21.2 729.6 5.4 4.4 90.6 15.5 45.7 1.9 18.4 11.8 16.5 12.9 3.0 4.8 1.6 4.7 6.1 35.9

1 70 000 130 415 23 500 197 446 3 960 041 40 440 40 000 728 427 142 572 354 258 8 600 133 035 59 580 130 830 67 000 29 536 21 477 13 282 19 127 36 975 309 084

106.5 194.5 155.7 107.4 184.2 134.3 109.0 124.4 109.0 129.1 217.4 138.0 197.5 125.9 192.4 101.9 221.2 123.5 245.2 164.4 116.2

36.6 2.4 9.7 12.9 3.8 1142.3

316 389 23 000 101 919 90 391 17 600 7 164 924

115.6 102.6 94.8 142.3 217.6 159.4

mgd, million gallons per day; gpcd, gallons per capita per day. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

(single family, commercial, industrial, institutional), (3) meter size, (4) meter readings and the dates of meter readings, (5) water use between meter readings, and (6) billing information (charges incurred, dates paid, etc.). The customer billing system is usually computerized and individual customer accounts can be sorted by customer type, geographical area (e.g., pressure zone), and other characteristics. Finally, water sales records are summaries of the individual water billing records. The sales data are aggregated by the billing cycle (i.e., monthly, bimonthly, semiannually, or annually) and by customer type. They show how much water is being sold to different types of customers but they do not show for what specific purposes the water is being used.

1.10.2.2.2 Industrial and commercial sector data Data on self-supplied commercial and industrial use within a given geographical area are available in areas where annual or monthly reporting of water is practiced. The industrial subsector includes water used for ‘‘industrial purposes such as fabrication, processing, washing, and cooling, and includes

Predicting Future Demands for Water Table 2 Estimates of self-supplied industrial and commercial water demand in 11 counties in Northeastern Illinois

Table 3

Water use in a sample of large power plants in Illinois

Plant name County

Boone Cook DeKalb DuPage Grundy Kane Kankakee Kendall Lake McHenry Will Total/ave.

Self-supplied withdrawal in 2005 (mgd)

0.57 123.73 2.54 0.96 6.99 4.34 5.09 0.78 13.88 6.58 24.97 190.43

Employment in self-supplied establishments

1 200 22 364 4 025 11 024 656 6 329 157 5 229 19 495 8 515 13 727 92 721

Unit selfsupplied withdrawals per employee (gped) 475.0 5 532.6 631.1 87.1 10 655.5 685.7 32 420.4 149.2 712.0 772.8 1 819.0 2 053.8

mgd, million gallons per day; gped, gallons per employee per day. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

such industries as steel, chemical and allied products, paper and allied products, mining, and petroleum refining,’’ and the commercial subsector includes water used for ‘‘motels, hotels, restaurants, office buildings, other commercial facilities, and institutions’’ (Avery, 1999). For a given geographical area such as a county or a hydrologic basin, industrial and commercial water withdrawals will depend on the number, type, and size of water users. Table 2 shows an example of self-supplied commercial and industrial withdrawal data for 11 counties in Northeastern Illinois.

1.10.2.2.3 Power generation sector data In the US, water withdrawn by power plants is classified as thermoelectric generation water use. It represents the water applied in the production of heat-generated electric power. The main use of water at power plants is for cooling. Nearly 90% of electricity in the United States is produced with thermally driven, water-cooled generation systems which require large amounts of cooling water (Dziegielewski and Bik, 2006). The three major types of thermoelectric plants include: conventional steam, nuclear steam, and internal combustion plants. In conventional steam and nuclear steam power plants, water is used primarily for cooling and condensing steam after it leaves the turbine. In this type of generation, the use of cooling water is essential because the collapse of steam volume in the condenser creates a vacuum, which affects the rotation of the turbine. Because the level of the vacuum depends on the removal of waste heat by cooling water, the cooling system is an integral part of the power generation process. Precise estimates of thermoelectric water use are difficult to obtain. The only consistent source of thermoelectric water-use data is the annual survey of power plants by the US Energy

167

2005 Water withdrawals (mgd)

Once-through flow plants Crawford Plant 503.3 Fisk Street Plant 222.2 Dresden Nuclear 415.6 Planta Waukegan Plant 758.6 Joliet 29 Plant 942.6 Joliet 9 Plant 415.3 Will County/ 917.9 Romeoville Plant Clinton Plant 810.4 Dallman Plant 328.1 Lakeside Plant 43.2 All once-through 5357.2 plants

2005 Gross generation (MWh yr1)

2005 Rate of withdrawals (gal. kWh1)

3 201 844 1 603 949 14 031 125

57.4 50.6 10.8

4 909 907 5 767 994 1 922 330 5 658 996

56.4 59.6 78.9 59.2

9 014 690 2 328 492 229 855 48 669 182

32.8 51.4 68.6 40.2

Closed-loop makeup water plants Vermilion Plant Powerton Plant Braidwood Nuclear Plant Kendall Co. Gen. Facility All closed-loop plants

2.8 25.9 49.8

702 950 10 120 133 20 390 274

1.43 0.93 0.89

2.5

1 367 008

0.67

80.9

32 580 365

0.91

a

Dresden plant uses a combination of once-through and pond recirculation system. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008 and WHPA and Dziegielewski (2008).

Information Agency (EIA). The resultant EIA-767 database consists of a series of data tables that present data on different aspects of the power plant operation (Table 3). The EIA-767 data tables are the main data sources for the analysis of thermoelectric water withdrawals (intake) and consumptive use. However, since 2006 the EIA discontinued collection of operational data from power plants; therefore, the more current data on water withdrawals have to be obtained directly from individual plants.

1.10.2.2.4 Irrigation sector data The irrigation sector includes self-supplied withdrawals of water for irrigation of cropland, turfgrass-sod farms, and golf courses. In the existing inventories of water use, the designation of irrigation water withdrawals includes ‘‘all water artificially applied to farm and horticultural crops as well as selfsupplied water withdrawal to irrigate public and private golf courses’’ (Solley et al., 1998). Irrigation water use is rarely measured and the reported data on water withdrawals are based on the inventory of the total acreage of irrigated area. The data on irrigated land are collected and reported by the US Department of Agriculture (USDA, 2009).

168

Predicting Future Demands for Water

1.10.3 Water-Demand Relationships

water. That is, in general, for k inputs,

1.10.3.1 Theoretical Models of Water Demand Guidance for empirical modeling of water demand can be derived from economic theory. From an economic perspective, water is considered to be a commodity (or an economic good) and it can be conceived as a final good to consume or as an input to the production of some other good or service. Demand for water therefore can be a final demand if the user is a consumer or a derived demand if the user is a producer whose demand is driven by the demand for other goods produced through the use of water. Theoretically, industrial and most of the commercial and public uses can be viewed as derived demands by producers, while the residential component of water demand can be viewed as a final demand by consumers (Hanemann, 1998). Different economic theories apply to these two economic categories of demand. The derived demand is described by the economic theory of production. The final demand can be described by the economic theory of consumer demand. Good examples of a complete mathematical treatment of the application of these two theories to water demand are provided by Hanemann (1998) and Renzetti (2002).

1.10.3.1.1 Derived demand of producers According to Hanemann (1998) the derived demand for water by a firm can be represented by four different types of relationships. These include both the long-run and short-run versions of conditional and unconditional demand functions. The fixed-output conditional demand function reflects optimization of the amounts of inputs (which include water) to produce a given level of output. The variable-output or unconditional demand function determines how the firm should select the level of output together with the corresponding inputs. The other two functions introduce the distinction between the short-run and long-run input demand functions. In a short run some inputs are assumed fixed, whereas in a long run all inputs are variable. The long-term conditional demand functions can be obtained by combining the production function (which represents production technology) with the firm’s behavior (i.e., cost minimization or profit maximization). The derivation of conditional or unconditional demand functions would start with an explicit formula for the production function and then solve the optimization problem. Two production functions which were frequently used in the past include the Cobb–Douglas and the constant elasticity of substitution (CES) formulas. A newer approach uses the duality relationship between a production function and the cost or profit functions. Hanemann (1998) used the duality approach to show the two demand functions – conditional and unconditional demand. The conditional function can be represented by the derivatives of the associated cost function or profit function with respect to the price of the input of interest. Accordingly, the conditional water-demand function can be written as a derivative of the cost function with respect to the price of

xk ¼ gk ðo1 ; y; on ; yÞ ¼

q Cðo1 ; y; oN ; yÞ ; k ¼ 1; y; N q ok ð1Þ

and the unconditional demand function is equal to minus the derivative of the profit function with respect to the price of water:

xk ¼ hk ðo1 ; y; oN ; pÞ ¼ 

q pðo1 ; y; oN ; pÞ ; k ¼ 1; y; N q ok ð2Þ

where C denotes the firm’s total cost of production; ok the price of the kth input xk, with k ¼ 1,y, N; y the volume of output produced per unit of time; p the profit; and p the price of output. There are many possible formulas for cost or profit functions that can be used with this approach including production functions which are less restrictive than the Cobb– Douglas or CES functions with respect to their implications for complementarity or substitution among inputs. As an example, Hanemann (1998) provides an illustration of the application of the derivatives in Equations (1) and (2) using the translog cost function:

lnC ¼ lnb0 þ by lny þ þ

X

X

ln ok þ 1=2dyy ðlnyÞ2

k

dky ln ok lny þ

k

1X dik lnoi lnok 2 ik

ð3Þ

where b‘s and d‘s are coefficients to be estimated with dik ¼ dki. The application of Equation (1) in combination with Equation (3) leads to a relatively simple share equation which represents the share of total cost devoted to each input:

ok gk ðo1 ; :::; oN ; yÞ C ¼ bk þ by ln y þ dky ln y þ dkk ln ok X þ oki ln oi

sk ðo1 ; :::; oN ; yÞ 

ð4Þ

ia k

The theoretical relationships given above provide some guidance for the development of empirical water-demand equations for industrial and commercial sectors. The main points are: (1) derived demand for water depends on the price of water (own elasticity) as well as the prices of all other inputs (cross-price elasticity), and (2) demand also depends on either the quantity or prices of the outputs. These theoretical relationships should be incorporated into econometric models of industrial water demand. However, there are some practical difficulties which limit a strict adherence to the economic theory of production. One limitation is the level of data aggregation. Water-use data are often aggregated for an entire industrial sector while the production and optimization relationships need to be estimated separately for specific industries based on the firm-level data. Other difficulties relate to (1) specification of the price of water, (2)

Predicting Future Demands for Water

availability of prices for inputs other than water, and (3) specification of the mathematical form of the demand function. Examples of empirical relationships which were obtained in studies of industrial demands can be found in the work by Renzetti (1992) and Dupond and Renzetti (1998).

1.10.3.1.2 Final demand of consumers Economic theory of consumer demand provides a basis for deriving relationships for final demand for water. The theory is based on the concept of the consumer’s utility function and the optimizing behavior of a consumer faced with a limited budget. Utility functions can take many forms; examples include the Cobb–Douglas function and its variant, the Stone– Geary function (Deaton and Muellbauer, 1999). For example, Hanemann (1998) derived a water-demand function based on the Stone–Geary function for the case of two goods (i.e., water and all other goods) of the form

x1 ¼ ð1  a1 Þg1 þ a1

y p2  a1 g2 p1 p1

ð5Þ

where x1 is the quantity of water consumed, y the consumer’s budget, g1 and g2 the minimum consumption levels (which specify the assumption that the consumer derives utility from a commodity only if the consumption exceeds gi), p1 and p2 the prices of the two goods, and a1 the exponent of x1 in the original utility function. The own price, cross-price, and income elasticities can be obtained by taking derivatives of this demand function. Generally, single-demand equations used in empirical studies of residential water demand (even if they include price of water, income, and price of other goods) are formally inconsistent with economic theory because their functional forms (i.e., model specification) cannot be derived from the maximization of the utility function (Hanemann, 1998). The correct specification of a linear demand function with these three variables would be

x1 ¼ a1  b

p1 y þg p2 p2

ð6Þ

Equation (6) indicates that demand for water depends on the relative prices and relative income and not on their absolute values which are typically used in the empirical equations which are found in the literature. Another issue in modeling final demand is related to the choice of explanatory variables to be included in the demand function along with the economic variables of price and income. Variables such as those describing the climatic conditions, demographic characteristics, or physical settings need to be included in the equation because they also affect the consumer’s utility. Different values of these variables would cause different levels of utility from the same level of consumption. Hanemann (1998) suggests that other variables should be introduced into demand functions by making one or more of the coefficients in Equation (6) a function of those variables. In summary, the theoretical relationships described above provide some guidance for the development of empirical

169

water-demand equations for residential sectors. The main points are as follows: 1. Final demand for water depends on the price of water (own elasticity) as well as the prices of all other goods (crossprice elasticity) and the level of income. 2. Demand also depends on factors other than price and income and these factors have to be incorporated into the demand function in conformity with economic theory. However, a strict adherence of water-demand studies to economic theory is rarely achieved. Equations (1)–(6) represent demand functions by individual producers or consumers for specific uses of water. Because the empirical data sets are usually aggregated over individual users and often combine different uses of water, Hanemann (1998) observes that empirical studies of existing data make some leap from a theory that applies to individual agents to more aggregate data. For example, a number of empirical studies addressed aggregate water use in the residential sector. Some notable examples include Howe and Linaweaver (1967), Foster and Beattie (1979), Howe (1982), Nieswiadomy (1992), and Epsey et al. (1997). Most of these studies use model specifications which are consistent with economic theory but the theoretical demand models are applied to aggregate data.

1.10.3.2 Empirical Models of Water Use Empirical modeling of water demand consists of the search for variables that help explain water demand and the determination of their relationships to water quantities used. The results of previous studies contain important information about the principal explanatory variables and their mathematical relationship to water demand. Because researchers have defined water demand in many different ways, numerous empirical models appear in the literature. In this section, special emphasis is placed on the models of aggregate demands, which consider total demands of (1) a group of water users who use water for a similar set of purposes or (2) various often dissimilar users within a defined geographical area. Although this section does not present a comprehensive review of the entire literature of water-demand modeling, a sufficient number of studies are included to provide a representative sample of the approaches that have been employed to explore water demand.

1.10.3.2.1 Configuration of data sets In econometric studies, data on economic activities are collected at either micro- or macro-levels. Observations on individual households, families, or firms are referred to as microdata. Regional- or national-level accounts and observations of entire industries are called macrodata. In the analysis of water use, the corresponding types of data are often referred to as disaggregate and aggregate data. Levels of data aggregation by purpose of water use range from the most disaggregate level of end-uses (e.g., toilet flushing or cooling tower makeup) to aggregated sector-wide totals, including domestic, industrial, or other uses. Water use can also be aggregated by summing it over various time periods (day, month, or year) and geographical areas (townships, cities, counties, or states).

170

Predicting Future Demands for Water

For analytical purposes, observations of water use (and of the corresponding explanatory variables) can be obtained and organized in several ways. In mathematical terms, we can describe each data configuration by designating the water use of entity i during time period t as Qit. Depending on which type of arrangement is used, the following three types of data configurations can be distinguished:

•

• •

Time series data, Qit. Recorded or estimated water use of an individual water user, group of water users, or all uses within a defined geographical area, during each time period t in a time series, where i ¼ constant and t ¼ 1, 2, y, T. Cross-sectional data, Qit. Recorded or estimated water use of each individual water user, sector, or geographical area i during time period t, where i ¼ 1, 2, y, n, and t ¼ constant. Pooled time series and cross-sectional data, Qit. Recorded or estimated water use of each individual user, sector, or geographical area i, in each time period t, where i ¼ 1, 2, y, n and t ¼ 1, 2, y, T.

In time series data, observations of all variables in the data are taken at regular time intervals (e.g., daily, weekly, monthly, or annually). In cross-sectional data, observations are taken at one time (either a point in time or time interval) but for different entities (such as households, firms, sectors of water users, cities, counties, or states). Pooled data combine both the time series and cross-sectional observations to form a single data matrix. A special case of pooled data is known as panel (or longitudinal) data, which represent repeated surveys of the same cross-sectional sample at different periods of time. The above types of data configurations form an empirical basis for developing water-use relationships. The mathematical form (i.e., linear, multiplicative, and exponential) and the selection of the right-hand side (RHS) or independent (explanatory) variables depend on the type and aggregation of water demand represented by the left-hand side (LHS) or dependent variable. Depending on the purpose for which water-use estimations are to be used, different representations of the dependent variable may be employed. For example, in hydrologic studies of surface and groundwater resources, water use is usually represented as daily, monthly, or yearly withdrawals at a point such as a river intake or a groundwater well. Because the water withdrawn is typically used (or applied) over a larger land area, an equivalent hydrologic definition of water use would be the use of water within a defined geographical area (e.g., an urban area, a township, a county, or a river basin or subbasin) which is obtained from a single point of withdrawal. While the quantities of water withdrawn in time can be precisely measured at the withdrawal points, they cannot be modeled without an appropriate consideration of the nature of water demands for which the withdrawals are made. The nature of water demand depends on the aggregation of water uses and users. The aggregation levels can range from a single user at a point location to a diverse group of users within a geographical area. For an individual user, water use can be represented as a sum of quantities of water used for specific purposes (or enduses) as

Qjt ¼

X i

qijt

ð7Þ

where Qjt is the total water use of an individual user j during time period t and qijt the water used for a specific purpose of use i, such as garden watering, washing, or cooling during time period t. At a higher level of aggregation, water use within a larger geographical area such as a city, county, or river basin can be represented as a sum of water use for several groups of users within a number of subareas:

Qt ¼

XXX j

k

Qjkgt

ð8Þ

g

where Qt is the aggregate water use of all individual users j within k user sectors within geographical subareas g and Qjkgt designates water use by individual users in the area. In the case of an urban water-supply system, j would represent individual customers, k would represent major user sectors such as the single-family residential or commercial sectors, and g would represent sections of the city or pressure zones. Generally, water use at any level of aggregation Qt can be modeled as a function of explanatory variables Xi. However, because different components of aggregate water demand may be determined by different sets of explanatory variables and different functional forms, more precise models can be obtained by disaggregating demand Qt into its components, especially sectoral demands and modeling each component separately.

1.10.3.2.2 Functional forms and model parameters The most common approach to modeling water demand relies on multiple regression techniques. Usually, the dependent variable is assumed to be a linear function of several independent variables. For example, if there are three independent variables, the model can be written as

Q ¼ a þ b1 X1 þ b2 X2 þ b3 X3 þ e

ð9Þ

where a, b1, b2, b3 are the estimated regression coefficients, X1, X2, X3 the independent variables assumed to affect independent variable q, and e the random error term. The coefficients of a and bi are estimated by finding the values that minimize the sum of the squared deviations for the observed values of the dependent variable from the values of the dependent variable predicted by the regression equation. In order for the ordinary least-squares (OLS) regression analysis to yield valid results, the error term is assumed to be normally distributed and to have zero mean, common variance across all observations and to be independent of all explanatory variables. Also, there should be no correlation between independent variables and the distribution of the dependent variable should be approximately normal. Two additional conditions are that (1) none of the independent variables can be an exact multiple (or linear) combination of any other independent variable and (2) the number of observations must exceed the number of coefficients being estimated. When the five basic assumptions of regression model are satisfied, the OLS procedure would produce unbiased estimates of the regression coefficients a and bi , which have minimum variance among all unbiased estimates.

Predicting Future Demands for Water

Alternatives to the linear model include the log–log model (which is linear in the logarithms of the dependent and independent variables) and the semi-log models (in which either only the dependent variable is or only the independent variables are transformed into logarithms). For example, the log-linear (or double log) model with k explanatory variables can be written as

Qit ¼ ea0

Y

Xit bk  eeit

X

Z¼

bk ln Xkit þ eit

ð10bÞ

k

In some empirical models, Qit is converted into per capita withdrawals, for example, public-supply withdrawals are divided by the population served in the study area. Equation (10a) can be used to represent an aggregate demand curve or a market demand curve. According to a priori (theory-based) expectations, the demand curve would be negatively sloped. The shape and position of the demand curve are determined by the values of other explanatory variables. For residential demand, these variables may include income, household size, temperature, and rainfall. According to expectations, the effect of increasing income would be to shift the curve to the right, so that the same price would result in progressively larger quantities of water being used. The effect of increasing precipitation during the growing season would shift the curve to the left. The entire demand curve does not have to be known in order to determine the effect of independent variables on water demand (Boland et al., 1984). It is usually sufficient to know how specified incremental changes in explanatory variables will affect water use. In the case of price, this information is contained in the slope of demand curve. The slope gives the incremental change in water use for an incremental change in price, at some position on the curve. Because of the units chosen for axes of the demand curve (dollar per unit of water use and units of water use), the slope of the curve has an inconvenient dimension (dollars per unit of water use squared) (Boland et al., 1984). It is customary, therefore, to use a dimensionless measure of the relationship, calculated by dividing fractional (instead of incremental) change in water use by fractional change in price. This dimensionless measure is known as elasticity. For the price variable, it is called the price elasticity of water demand. It is defined for an arc of the curve as

Q2  Q1 Q Z¼ P2  P1 P

dQ P dP Q

ð12Þ

where water use is a function of price and other variables, the ordinary derivative in Equation (12) is replaced with a partial derivative:

Z¼

where variables and coefficients are as defined in Equation (9). By taking the natural logarithm of both sides of Equation (10a), the following log-linear model is obtained:

ln Qit ¼ a0 þ

specific point on the curve as follows:

ð10aÞ

k

qQ P qP Q

ð13Þ

Both arc and point definitions give a dimensionless elasticity, which is expected to be negative (because the demand curve is negatively sloped). Price elasticity may be interpreted as the percentage change in quantity which would result from a 1% change in price. A price elasticity of  0.3, therefore, indicates that 1.0% increase in price would be expected to result in a 0.3% decrease in quantity demanded (use). Conversely, a 1.0% decrease in price would produce a 0.3% increase in quantity demanded. Depending on the magnitude of the calculated elasticity, the demand is said to be perfectly inelastic when Z ¼ 0.0; relatively inelastic when 0.0 4 Z 4 1.0; unitary elastic when Z ¼  1.0; relatively elastic when  1.0 4 Z 4 –N; and perfectly elastic when Z ¼ –N. In other words, demand is said to be relatively inelastic when quantity changes less than proportionately with price, and relatively elastic when quantity changes more than proportionately with price. Elasticity can be calculated for all other explanatory variables using the following formulas for the most common functional forms. For any variable X, the method of calculating the elasticity value from the estimated regression coefficients in four different functional forms is as follows: 1. For the linear form with no log transformation of any variables, the elasticity with respect to X is calculated in terms of the means of the variables as

Q ¼ a þ bX and Z ¼ b

X Q

ð14Þ

2. For the log–log form (also called double-log), the elasticity with respect to X is constant and equal to the regression coefficient of X:

lnQ ¼ a þ b ln X and Z ¼ b

ð15Þ

3. For the semilog form where the dependent variable is untransformed and the dependent variables are transformed into their natural logarithms, the elasticity with respect to X is

ð11Þ

where Q  ¼ ðQ1 þ Q2 Þ=2 and P  ¼ ðP1 þ P2 Þ=2. A more frequently used definition is based on the derivative of the demand function, and yields the elasticity at a

171

Q ¼ a þ b ln X and Z ¼

b X

ð16Þ

4. For a semilog form where only the dependent variable is transformed into logarithm, the elasticity with respect to X

172

Predicting Future Demands for Water Table 4 studies

is directly proportional to X:

lnQ ¼ a þ bX and Z ¼ bX

ð17Þ

To compare empirical studies of water demand, it is both simple and instructive to compare only the elasticities with respect to the explanatory variables, thus getting around the difficulty of interpreting direct comparison of regression coefficients. The following sections compare the results of past studies in terms of the elasticities of key explanatory variables.

Variables used in municipal and residential water-use

Explanatory variable

Variable definitions

Population

Number of users per account Population density Average number of residents per water meter Marginal price of the last unit of water used Average price or total water billed divided by total use Ratio of average price to marginal price Monthly income per capita per dwelling unit Residential property value Per capita income Average household income Median household income Percent of families in low income bracket Imputed rent derived from home value Number of housing units Percent of units by housing type Percent of units occupied by owners Average housing units per acre Number of rooms per dwelling Number of bathrooms House size Lawn size Outdoor irrigable area Median number of rooms Building age Number of persons per dwelling unit Age of the head of household or spouse Educational attainment of the head of household Median age of householders Percent in family households Percent of married households Summer average evapotranspiration rate Monthly effective evapotranspiration rate Monthly rainfall Average monthly rainfall between last spring freeze month and first fall freeze month Precipitation per billing period Mean annual rainfall Precipitation during growing seasons Average temperature for months between last spring freeze month and first fall freeze month Monthly average temperature Monthly average of maximum daily temperatures Number of days without significant rainfall (Z 0.04 in) times the month’s average temperature Number of retail establishments Value added in manufacturing Number of employees in all sectors Number of production workers in manufacturing

Water price

Income

1.10.3.2.3 Elasticities with respect to major determinants of water use Past empirical studies of water demand have used a broad array of possible explanatory variables. Table 4 lists variables in seven major categories and their definitions. These were found in empirical studies of municipal and residential water use. Estimated elasticities of explanatory variables can be obtained from the published empirical equations either as constant elasticities in double-log models or as calculated elasticity values at the mean values of the dependent and independent variables in linear and semilog models. The empirically derived elasticities of key explanatory are summarized in the following. Price elasticity. Economic theory assumes that consumers respond to economic incentives by adopting behaviors that maximize their well-being. In one of the earliest studies of urban water demand, Metcalf (1926) documented a relationship between water use and price that implied price elasticity of demand in the range of  0.40 to  0.65. A substantial body of literature has been published since to confirm that consumers respond to changes in the price of water (Boland et al., 1984; Epsey et al., 1997). Empirical estimates of the price response (elasticity) generally range between  0.1 and  0.9 with higher (absolute) values in industrial and agricultural uses. These values of price elasticity indicate that a 1.0% increase in price would result in a 0.1– 0.9% decrease in water use. Table 5 shows the range and most likely values of price elasticities of water demand for several types of water users. The price elasticity values were obtained from 60 empirical studies of water demands. While these elasticity coefficients indicate that demand is relatively inelastic with respect to price, significant increases in price are expected to result in major reductions in demand. During water shortages, rationing through pricing has proved to be an effective strategy for achieving significant reductions in demand. For example, during the 1988 water shortages in Santa Barbara, California, the price was raised to 27 times the normal level (from $1.09/100 cubic feet or $0.39/m3 to $29.43/100 cubic feet or $10.40/m3) to deter all but the most essential uses of water in the city (Ferguson and Whitney, 1993). As a result, the sector-wide demands were reduced by 56% in the single family, 41% in multifamily, and 20% in commercial sector. Also, average wastewater flows were reduced by 45%. Although price increases were accompanied by the implementation of a sprinkling ban and other conservation measures, the effect of the price increase on indoor

Housing

Family composition

Weather

Othera

a

Used in municipal demand models only.

use (unaffected by sprinkling restrictions) implies a short-term elasticity of  0.22. Income elasticity. Economic theory indicates that together with price, income is a key determinant of residential water demand, because the latter determines the consumer’s ability to pay for water. Consumers decide on what features and

Predicting Future Demands for Water Table 5

173

Empirical price elasticities of water demand

Demand category

No. of studies

No. of estimates

Range of price elasticitiesa

Median value

Combined urban demand Residential demand Single-family only Nonresidential demand Commercial Industrial Institutional Agricultural irrigation

25 58 24 15 6 19 3 10

93 256 94 160 53 101 54 34

– – – – – – – –

– – – – – – – –

0.11 0.18 0.22 0.27 0.24 0.33 0.24 0.24

to to to to to to to to

– – – – – – – –

0.58 0.50 0.48 0.87 0.92 0.88 0.94 0.97

0.40 0.33 0.31 0.54 0.34 0.58 0.47 0.46

a

The range shows the 25th and 75th percentile in the distribution of reported estimates.

Table 6

Empirical income elasticities of residential water demand

Demand category

No. of studies

No. of estimates

Range of income elasticitiesa

Median value

Residential demand Single-family only Multifamily only Municipal demand All sectors

37 24 2 23 86

137 82 2 38 259

0.20 0.10 0.22 0.19 0.14

0.37 0.18 0.22 0.31 0.31

to to to to to

– – – – –

0.61 0.39 0.23 0.58 0.55

a

The range shows the 25th and 75th percentile in the distribution of reported estimates.

conveniences they want in their residences, and what technology they want to employ to achieve them, considering the required investments and the price of water. Table 6 compares income elasticity estimates which were derived from 86 studies of residential (and combined municipal) water demand. For all studies, the range of 25th and 75th percentile estimates is between 0.14 and 0.55 with a median value of 0.31. Elasticities with respect to air temperature and precipitation. Weather conditions influence water demand because some uses of water such as landscape or crop irrigation are sensitive to variables such as precipitation, air temperature, or evapotranspiration. Table 7 shows elasticity estimates for air temperature and precipitation which were derived from 30 studies of residential and nonresidential water demand. For all studies, the range of 25th and 75th percentile estimates of the elasticity of temperature is between 0.43 and 2.15 with a median value of 1.15. The corresponding percentile values of the elasticities of precipitation are between  0.03 and  0.19 with a median value of  0.07. These elasticities indicate that on average demand is more than 10 times more responsive to changes in air temperature than to changes in precipitation. Elasticity with respect to production output. Output in manufacturing activities is often used as an explanatory variable in industrial water-use studies. However, different studies used different proxies for output, such as output value, number of employment hours, and number of employees. Table 8 lists 10 estimates of elasticities for output variables. All estimates are positive, and range from 0.48 to 1.94. Output elasticities 41 indicate that the use of water increases faster than output. The values in Table 8 show that this is the case for some industrial end-uses of water (i.e., cooling, processing, and steam generation) and some industrial categories. Elasticities of output for stone products, photographic

equipment, heavy industry, and paper production are o1, thus indicating that water use increases slower than output.

1.10.3.3 Other Water-Use Relationships The preceding sections described water-use relationships that derive from the economic theory of water demand. However, the limited availability of data on economic variables and the aggregate nature of data on water use often preclude the development of econometric models of water demand. A noneconomic approach is sometimes used to develop wateruse relationships which represent water requirements for different types of water users. Such approaches are explicitly or implicitly based on the assumption that the quantities of water used relate to a technical or physical requirement and are unaffected by economic choice. Nevertheless, the requirements approach remains an option for quantifying water use where econometric models cannot be developed. Two examples of the requirements models are described here.

1.10.3.3.1 Cooling water requirements In once-through cooling systems in steam-based thermoelectric power plants, theoretical water requirements are a function of the amount of waste heat that has to be removed in the process of condensing steam. According to Backus and Brown (1975), the amount of water for 1 megawatt (MW) of electric generation capacity can be calculated:

L¼

6823ð1  eÞ Te

ð18Þ

where L is the amount of water flow in gallons per minute (gpm) per MW of generating capacity; 6823 the units conversion factor; T the temperature rise of the cooling water in F;

174 Table 7

Predicting Future Demands for Water Empirical elasticities of two weather variables

Demand category

No. of studies

Air temperature Residential demand Single-family only Multifamily only Municipal demand Nonresidential All sectors Precipitation

7 6 1 5 2 21

43 11 1 16 3 74

Residential demand Single-family only Multifamily only Municipal demand Nonresidential All sectors

11 7 2 10 1 30

57 22 3 43 6 131

Range of income elasticitiesa

No. of estimates

0.44–3.58 0.88–2.00 0.35–0.35 0.53–1.58 0.02–0.81 0.43–2.15  0.04  0.01  0.04  0.05  0.01  0.03

to to to to to to

Median value

1.37 0.88 0.35 1.31 0.02 1.15

 0.24  0.09  0.12  0.21  0.15  0.19

 0.09  0.02  0.12  0.09  0.03  0.07

a

The range shows the 25th and 75th percentile in the distribution of reported estimates.

Table 8

Examples of elasticities of output variables in industrial water-use models

Study/author

Measure of output

Elasticity

Notes

De Rooy (1974)

Output value

Dziegielewski et al. (1990)

Total employment

Renzetti (1988)

Total number of employee hours

Renzetti (1993)

Output value

1.21 1.22 1.19 0.48 0.60 1.11 0.69 1.94 0.72 0.61

Cooling water use only Processing only Steam generation only SIC328 (cut stone and stone products) SIC386 (photographic equipment and supplies) SIC334 (secondary nonferrous metals) Heavy industry Light industry Self-supplied paper industry Self-supplied textile industry

and e the thermodynamic efficiency of the power plant, expressed as decimal fraction. For example, in a coal-fired plant with thermal efficiency e of 40% and the condenser temperature rise of 20F, the water flow rate obtained from Equation (18) would be 512 gpm per MW. For a typical 650 MW plant, operating at 90% of capacity, the theoretical flow rate would be nearly 300 000 gpm, or 431.3 mgd. The daily volume of cooling water is equivalent to approximately 31 gallons per 1 kWh of generation. According to Croley et al. (1975), in recirculating systems with cooling towers, theoretical makeup water requirements are determined using the following relationship:

W ¼E

1 1

c c0

ð19Þ

where c/c0 is the concentration ratio which compares the concentration of solids in makeup water to their concentration in the recirculating cooling water and E the evaporative water loss which, for a typical mean water temperature of 80 F, can be calculated as

E ¼ ð1:91145  10 6 Þ  aQ

ð20Þ

where a is the fraction of heat dissipated as latent heat of evaporation (for evaporative towers a ¼ 75  85%); 1.911 45  106 the units conversion factor; and Q the rate of heat rejection by the plant in Btu h1, which can be calculated as

Q ¼ 3414426  P 

1e e

ð21Þ

where P is the rated capacity of the plant in MW; 3 414 426 the units conversion factor; and e the thermodynamic efficiency of plant expressed as a fraction. Again, for a typical 650 MW coal-fired plant with 40% efficiency, the heat rejection would be 3329 million Btu h1 and the evaporative water loss would be 5091 gpm. At the concentration ratio c/c0 of 0.25, the makeup water flow would be 6788 gpm or 0.63 gallons per 1 kWh of generation.

1.10.3.3.2 Supplemental irrigation water requirements Water required for supplemental irrigation depends on soil moisture deficit during the growing season which can be derived based on rainfall data. The total seasonal application depth can be determined according the method developed by the Illinois State Water Survey. The method is based on weekly precipitation records for the growing season from 1 May to 31

Predicting Future Demands for Water

August (Dziegielewski and Chowdhury, 2008). Rainfall deficit is calculated by accumulating weekly deficits or surpluses over the consecutive weeks of the growing season. If more than 1.25 inches (in) of rain falls during the first week of the growing season, one-half the amount of rain exceeding 1.25 in is added to the rain amount during the following week. If o1.25 in of rain falls during the first week, the difference between the actual rainfall and 1.25 in is the rainfall deficit that is assumed to be the quantity of water (in inches) applied by irrigation that week. For each subsequent week during the growing season, one-half of the cumulative rainfall during the previous week in excess of 1.25 in is added to the rainfall amount for the week. If the cumulative rainfall amount for a week is less than 1.25 in, then the difference between the actual rainfall and 1.25 in is the rainfall deficit that is assumed to be the quantity of water (in inches) applied by irrigation that week. The rainfall deficits for each week are then added to determine the total irrigation water use during the growing season. This procedure can be expressed in mathematical terms as follows: 1. If the total rainfall in the first week r1o1.25 in, then

d1 ¼ r1  1:25

ð22Þ

2. If the total rainfall in the first week r141.25 in, then

r2

e

d1 ¼ 0 ¼ r2 þ ðr1  1:25Þ=2 d2 ¼ r2 e  1:25

ð23Þ

where r2 e is the effective rainfall in week 2. In week 2, again, the precipitation deficit will be zero if r2 e 41.25 in, and one-half of the precipitation surplus will carry to the next week. The total seasonal rainfall deficit for the 18 weeks (i.e., 4 months) which make up the irrigation season is calculated as

dt ¼

18 X

di

ð24Þ

i¼1

Here, the values of precipitation deficit represent the total depth of water application in inches during the growing season. Thus, the requirements for supplemental irrigation water can be determined using the following formula:

Qt ¼

325; 851 At  dt 12  365

ð25Þ

where Qt is the annual (seasonal) volume of irrigation water in mgd in year t; At the irrigated land area in acres in year t; dt the depth of water application in inches in year t; and the conversion factors represent: 325 851 gallons/acre-foot, 12 inches/foot, and 365 days/year.

175

1.10.4 Demand Forecasting Techniques 1.10.4.1 Forecasting Principles and Criteria A basis for forecasting future quantities of water demanded is required if any type of planning is to be undertaken. As Gardiner and Herrington (1986) simply state: ‘‘y planning, of virtually any kind, requires forecasting’’ (p. 7). In planning for sustainable future water supply, forecasts of water demand which include predictions of improvements in efficiency of water use form a basis of long-term plans for balancing water demand with supply. Other planning activities which require forecasts of water demand include expansion of the capacity of water-supply infrastructure, allocation of limited water supplies among different users, as well as short-term operational and financial planning. Boland (1998) provides an excellent exposition on the basic premises and principles of forecasting. He defines a forecast as a statement about the future. In his earlier writings, he used a more detailed definition which described a forecast as a conditional statement about the future which is likely to materialize if the forecasting assumptions are proved to be correct (Boland, 1985). Boland (1998) also distinguishes the term forecast from the related terms prediction, projection, and extrapolation in terms of implied method or procedure for preparing the statement about the future. Thus, prediction implies nothing about the method and could be an entirely subjective and judgmental statement; extrapolation represents a continuation of past trends and projection suggests a prediction which is influenced (indirectly) by past trends (Boland, 1998: 81). Recently, the International Panel on Climate Change (IPCC) considered some clarifications to this terminology and appears to prefer an alternative term scenario defined as ‘‘ya coherent, internally consistent and plausible description of a possible future state of the world. It is not a forecast; rather, each scenario is one alternative image of how the future can unfold’’ (IPCC, 2008). In a recent study of regional water demands in Northeastern Illinois (Dziegielewski and Chowdhury, 2008), the IPCC definition was adopted in describing water-demand scenarios. The adopted definition ensured that the scenarios would not represent most likely forecasts or predictions, nor would they set upper and lower bounds of future water use but instead they would only describe three alternative paths in demand growth because different assumptions or conditions could result in water demands that are within or outside of the range represented by the three scenarios. In essence, a forecast or a scenario is a translation of a set of assumptions into a future outcome (i.e., a quantity of water used at some point in time). These assumptions are the basis and the main component of any forecast. Other forecast components such as its structure (i.e., time step or level of disaggregation) or computational algorithms and empirically derived models are important but not as critical as the forecasting assumptions. Forecasting assumptions are a part of each forecasting method, regardless of the level of its sophistication. Typically, the simpler methods have only a few assumptions which are likely to be crude and difficult to verify. More elaborate methods, such as the IWR-MAIN model (to be discussed later in this chapter), rely on hundreds of assumptions in deriving a

176

Predicting Future Demands for Water

forecast. Although it is possible to document and make all such assumptions explicit, it is rarely done. Only some of the assumptions, usually those judged by the analyst to be important, are made explicit. However, because the accuracy of the forecast cannot be assessed until the forecast period has passed, its putative validity can be established only by assessing the plausibility of forecast assumptions. Boland (1998) discusses the objective and subjective components of the forecast in the context of a two-step process which consist of explanation and prediction. The explanation step involves the analyses of historical water use and represents the objective part of the forecast. The prediction step applies the factors and relationships which explained water use in the past to generate a forecast. This step necessarily reflects the subjective judgment of the analyst. This subjective judgment is present in all the forecasting assumptions mentioned above. Therefore, a forecast should attempt to make a clear and credible portrayal of the determinants and assumptions behind future water demand. In order to increase acceptability of the forecast by decision makers as well as by other analysts, while preparing the forecast it is important to ensure that: 1. historical water-use data are presented and analyzed for trends and underlying causes and relationships; 2. historical trends and causes are differentiated across user sectors and geographical parts of the study area; 3. major factors influencing water usage rates are considered (e.g., prices, income, and housing densities), and the estimated models are correctly specified; 4. all assumptions are explicit and supported by analysis of past trends or a consensus on future trends; and 5. forecasts utilize an official or a consensus forecast of population and economic growth data. The above elements of the forecast can help ensure that the forecast is understood and accepted by decision makers. The following section describes the specific analytical methods which can be used in constructing a forecast.

1.10.4.2 Forecasting Models and Procedures All forecasts attempt to predict the future value of water use, Qt, as a function of one or more explanatory variables and associated assumptions about the forecasting method and related parameters. The methods differ in terms of the number of explanatory variables and the form of the functional relationship. The forecasting methods also differ with respect to the structure of the forecast, especially in terms of separation of demands into more homogeneous categories of water use. This section describes a range of methods which can be found among the past forecasts of water demand.

In fitting trend lines, several alternatives for functional form exist, including linear, exponential, and logarithmic. When a linear trend line is selected, the future value of water use would be calculated as

Qf ¼ Qt þ bðf  tÞ

ð27Þ

where Qt is water use during the base year t (or the last year with known water demand), b the annual increment in water use Q, and f a future year of the forecast. In an exponential model where the annual (fractional) percentage change r is (b  1)100% when b40, the exponential growth of the future water use is calculated as

Qf ¼ Qt ð1 þ rÞðf tÞ

ð28Þ

This is a well-known equation for compounding interest in financial calculations. Finally, in a logarithmic (log) model, future water use would be increasing at a decreasing rate. A variant of the logarithmic function is a form linear in its logarithms. Using the multiplicative form the log-linear equation, the future water use can be calculated as

Qf ¼ Qt

 b f t

ð29Þ

where t is the time period of the forecast assuming (t – n) ¼ 1 for the first time period of the historical data series. In this trend function, the percent rate of growth is inversely proportional to time. The trend extrapolation method is often used with aggregate data. The greatest difficulty of this method is deciding on what type of function represents the best fit to the historical data. Both linear and nonlinear trends (e.g., exponential or logarithmic) can show equally good fit to the historical data, but their extrapolation into the future may produce considerably different results. Another significant problem is the main forecasting assumption behind this method: that the historical rate of growth will continue into the future and produce the same effects. Boland (1998) considers this method to be ‘‘yto simplistic for virtually any application’’ (p. 85) because of its implicit assumption that water use is explained by the passage of time and that all other variables which are known to affect water use are perfectly correlated with time or their effects cancel each other. However, the use of time trend as one of the explanatory variables in water-use models can be helpful in capturing the residual effects of unspecified variables once the effect of known factors which affect water use is accounted for. Trend extrapolation is also used in deriving future values of some explanatory variables.

1.10.4.2.1 Time trend forecasting 1.10.4.2.2 Water requirement forecasts

Future water demand can be determined by extending the historical trend in the past records of water use. This method would be termed extrapolation as defined by Boland (1998). The only analytical step in this method is finding the functional form g in the equation

A longstanding forecasting practice is to assume that water use is proportional to the size of a water-using activity or the number of water users. This proportionality can be expressed using a linear model

Qf ¼ gðQt ; Qt1 ; :::; Qtn Þ

Q ¼ a þ bN

ð26Þ

ð30Þ

Predicting Future Demands for Water 250

or a multiplicative (log-linear) model

ð31Þ

ð32Þ

The assumption of strict proportionality is often criticized by the economists because it ignores other factors that can affect water use, and in its application to derived demands it treats the size variable which is often the production output or employment as exogenous to the firm’s decision on water use. However, the representation of water demands as the number of users (or a driver of water use) times average rate of usage (or intensity of water use) is a convenient and practical forecasting approach. A well-known example of this approach is the per capita requirements method, which has been widely used in forecasting urban water use. According to this method, future values of the volume of publicly supplied water for a city or municipality are often obtained by multiplying future population (which represents N in Equation (32)) with an assumed per capita rate of water use (as b in Equation (32)) This method was widely used in the past because the metric of per capita water use can be easily obtained from the production records (see Section 1.10.2.2.1) and total population served can be obtained from the population census data. However, the validity of this method depends on the constancy of the per capita use rate over time. In most urban water-supply systems, per capita use rate is not constant; it fluctuates from year to year and often exhibits long-term increasing or declining trends. For example, Figure 1 illustrates the long-term changes in per capita water use in New York City. The per capita rate had been increasing from 101 gpcd in 1915 to 208 gpcd in 1988 but had declined to 134 gpcd by 2006. A per capita forecast prepared in any year in the past would not have captured the changes in per capita rates of use. Because the per capita approach relies on the assumption that the per capita rate will remain constant throughout the forecast period, the validity of per capita forecasts is doubtful. The main problem is that it ignores the changes in the structure and composition of urban demands over time and also ignores the effects of future changes in the determinants of different component demands. However, the unit-use coefficient methods can have some validity when applied to disaggregate demands of more homogeneous sectors or categories of water use. Examples of disaggregate requirement models are given in the following. Disaggregate requirement forecasts. When total demand in a geographical area is disaggregated by sectors, demand in sector k can be represented as a product of the number of water users

100 50 0 00 19 10 19 20 19 30 19 40 19 50 19 60 19 70 19 80 19 90 20 00 20 10

Q ¼ bN

150

19

where a, b, b, and g are treated as constants at a given level of aggregate demand. Although this simple approach is criticized by economists (Hanemann, 1998), it is widely used in forecasting water demands for industrial, commercial, and agricultural categories. Aggregate requirement forecasts. By forcing the intercept in Equation (30) to be zero or assuming that the exponent g in Equation (31) is 1, we obtain a simple requirements model in which total water use is strictly proportional to N:

200

GPCD

Q ¼ bN g

177

Year Figure 1 Historical per capita rates of water use in New York City.

and the average rate of use within the sector:

Qkt ¼ S Nkt  qkt k

ð33Þ

where Nkt represents the number of users (or other units) in sector k at time t and qkt the unit use coefficient (or average rate of water use per user) in that sector. Similarly, the total demand within the geographical area can be represented as

Qt ¼ S S S Nkgjt  qkgjt k g j

ð34Þ

where k denotes the disaggregation of water use into homogeneous sectors of water users (e.g., residential, commercial, industrial, and institutional) and g represents the spatial disaggregation of water use into various geographical subareas that are relevant for planning purposes. An example of a unituse coefficient qkgjt could be average use in single-family buildings of the residential sector in a suburban section of a city. This single-coefficient model can be extended by expressing the average rates of water use within each sector as a function of one or more explanatory variables. Usually, the dependent variable is assumed to be a linear function of more than one independent variable. For example, if there are two independent variables, the theoretical model is similar to Equation (8)and can be written as

q ¼ a þ b1 X1 þ b2 X2 þ e

ð35Þ

where a, b1, b2 are estimated regression coefficients, X1, X2 the independent variables assumed to affect independent variable q, and, e the random error term. There are also forecasting methods that are based on landuse categories. Such approaches typically represent a particular case of single coefficient requirement models. The land-usebased models display data in a way that is convenient for infrastructure planning and city master planning. For example, DCSE, a California-based software product and consulting firm, has developed a geographic information system (GIS)based water-demand forecasting procedure for estimating

178

Predicting Future Demands for Water

water demands in response to changes in land-use and related use factors. Water-demand projections are based on distribution of land-use categories and the corresponding water-use factors. Because land use in a service area can change due to conversion of land to residential, commercial, industrial, and other urban uses, the result will be a change in total water use.

1.10.4.2.3 Demand forecasts In practice, these methods require that regression models are constructed using the proper specification of the economic variables such as price and income and the level of output in the case of derived demand (see Section 1.10.3.1). It should be noted that econometric models are derived from observations which represent the points of intersection of demand and supply curves, and they usually represent reduced forms, as opposed to the structural equations which represent the true demand functions. Unfortunately, demand models are generally available only for single-family residential water use. Studies of multifamily residential sector and major industrial, commercial, and agricultural sectors are very limited in number. Nevertheless, forecasters should consider developing econometric models in order to probe the validity of estimated (or assumed) wateruse relationships.

1.10.4.3 Dealing with Forecast Uncertainty As mentioned earlier, all forecasts of future water demand are inherently uncertain. Generally, the uncertainty associated with the analytically derived future values of water demand can come from a combination of the following distinct sources: 1. Random error. The random nature of the additive error process in a linear (or log-linear) regression model which is estimated based on historical data guarantees that future estimates will deviate from true values even if the model was specified correctly and its parameter values (i.e., regression coefficients) were known with certainty. 2. Error in model parameters. The process of estimating the regression coefficients introduces error because estimated parameter values are random variables which may deviate from the true values. 3. Specification error. Errors may be introduced because the model specification may not be an accurate representation of the true underlying relationship. 4. Scenario uncertainty. No future values of any model variables can be known with certainty. Various assumptions must be introduced when projections are made for the water-demand drivers (e.g., population, employment or irrigated acreage, income, price, precipitation, and other explanatory variables). The first three sources of error can be addressed by a careful analysis of the data and model parameters. The fourth source of error – the scenario error (or assumption error) – requires an explicit evaluation of assumptions through sensitivity analysis, Monte Carlo analysis, or the use of scenario-based forecasts.

1.10.4.3.1 Model-dependent prediction intervals In econometric forecasts, each empirically derived model can be tested for specification error by using Ramsey’s specification tests (Ramsey (1969), the Breusch–Pagan–Godfrey test Breusch and Pagan (1979)), Glejser’s test (Glejser, 1969), and Harvey’s test and White’s test (White, 1980) for heteroscedasticity. The specification and heteroscedasticity tests allow the analysts to develop predictive equations which minimize the errors from misspecification of the model and biases in model parameters. Other model-dependent errors can be quantified using confidence intervals (Dziegielewski et al., 2005). For example, assuming that the errors are normally distributed in a log-linear model in which Y designates water use, it can be shown that

EðYjexplanatory variablesÞ ¼ e se

2

=2

ðe ln Y Þ

ð36Þ

Thus, in log-linear models, the predicted value denoted as Y˜ is given by 2 Y˜ ¼ e s^e =2 ðe ln YÞ

ð37Þ

^e 2 is the mean square error of the log-linear model and where s lnY^it the predicted value obtained from the log-linear models. It is straightforward to obtain the in-sample prediction confidence intervals in a linear model. However, in a log-linear model, the in-sample prediction intervals are obtained under the assumption that the errors are normally distributed. Thus, for normally distributed errors the variance of Y˜ in Equation (37) is estimated by

  ˜ ¼ exp 2lnY þ Vii s ^2e VarðYÞ "  m=2  m #   ^2e s 2^ s2e 2 ^e 1   1  exp Vii s m m ð38Þ ^e 2 is the square of the standard error of the logawhere Vii s rithmic prediction (i.e., lnYit ), m the degrees of freedom, ^e 2 the mean square error of the log-linear model. The and s standard error of Y^ is denoted as

˜ ¼ SEðYÞ

qffiffiffiffiffiffiffiffiffiffiffiffiffiffi ˜ VarðYÞ

ð39Þ

0 Assuming that the Y˜ s are asymptotically normally distributed, the confidence interval for the prediction can be obtained as ˜ Y˜ þ za=2  SEðYÞ, ˜ where za=2 is the critical ½Y˜  za=2  SEðYÞ; value from a normal distribution for a prespecified a. However, for the out-of-sample predictions, the square of the ^e 2 ) is not standard error of the logarithmic prediction (i.e., Vii s available. To rectify this, one can use the average standard error of predictions (average over all observations in the historical data).

1.10.4.3.2 Dealing with forecast assumptions error The uncertainty caused by forecasting assumptions can be addressed by using scenario forecasts, sensitivity analysis, or probability forecasts. The scenario approach requires the

Predicting Future Demands for Water

development of scenario narratives and selection of a set of key forecasting variables and assumptions for each scenario. Usually, only a small number of scenarios are developed. The purpose of the scenarios is to capture future water demands under different sets of possible future conditions. Although the scenarios convey a sense of the range of future water use, they are not constructed to set upper and lower bounds of future values. Different sets of assumptions or conditions could result in demands that are within or outside of the range represented by the defined scenarios. Sensitivity analysis evaluates the sensitivity of the forecast values to changes in forecasting assumptions – one assumption at a time. In the case of the values of explanatory variables, alternative forecast values are generated by changing the value of the variable of interest while keeping the values of all other variables and forecasting assumptions unchanged. For key explanatory variables, the same information can be obtained by using the values of constant elasticities of water demand with respect to each variable and percent change in variable value. However, the effects of other assumptions are often less explicit and can be best determined by generating the alternative forecast values. Finally, probabilistic forecasts can be obtained by simultaneously varying two or more assumptions and generating a probability distribution of the values of water-demand forecast. Often, Monte Carlo simulation is used to generate very large numbers of outcomes based on distribution functions for the independent variables. Table 9 (IWR, 2001) shows an example of a probabilistic forecast for a water demand in public systems in a region in Virginia. The useful feature of such a forecast is the exceedance probability of the forecast values. The results in the table show that the mean forecast value for 2050 is 90.95 mgd and that there is only a 5% chance that the value of 96.94 mgd will be exceeded.

Table 9

179

Example of a probabilistic forecast

Year/water demand

2010

2050

Minimum Maximum Mean Std dev. Variance Skewness Kurtosis Distribution percentiles 5% 10% 15% 20% 25% 30% 35% 40% 45% 50% 55% 60% 65% 70% 75% 80% 85% 90% 95%

62.71 71.67 67.16 1.34 1.81 0.04 2.73

80.23 105.22 90.95 3.50 12.28 0.25 3.02

64.98 65.43 65.76 66.01 66.22 66.42 66.61 66.78 66.97 67.14 67.32 67.50 67.69 67.89 68.10 68.34 68.61 68.93 69.38

85.45 86.58 87.30 87.89 88.44 88.98 89.46 89.93 90.40 90.84 91.28 91.75 92.20 92.71 93.23 93.85 94.56 95.49 96.94

From IWR (Institute for Water Resources) (2001) An evaluation of the risk of water shortages in the Lower Peninsula, Virginia. Prepared by Werick WJ, Boland JJ, Gilbert J, Dziegielewski B, Kiefer J, Massmann J, and Palmer RN. IWR Special Report. Alexandria, VA: US Army Corps of Engineers.

1.10.4.4.1 End-use accounting system

1.10.4.4 Forecasts with Conservation The increasingly important role of demand-side options in water-supply planning creates needs for methods to estimate the effects of various water conservation programs on future water demands. However, the effects of demand-side programs cannot be assessed without a detailed knowledge of water uses in a study area and without an understanding of the important factors that influence them now and will influence them in the future. Thus, the most important feature of forecasting the impacts of water conservation is a high level of disaggregation of water demands. In order to estimate the effects of water conservation measures on future demand, it is usually necessary to disaggregate water demand into the specific end-uses of water. The conservation impacts (i.e., water savings) are usually estimated as the reduction in average rates of water use for specific purposes. For example, in the residential sector average household water use can be represented as a summation of average water used for toilets, showers, kitchen faucets, washing machines, and landscape irrigation. These end-uses may each be affected by water conservation measures which would result in a lower average per household use.

Forecasting methods which focus on impacts of long- and short-term demand management measures usually employ a highly disaggregated end-use accounting approach. Dziegielewski et al. (1993) proposed the following equation for estimating water use for each end-use:

" qe ¼

X

! Mu Su

# UþK F A

ð40Þ

u

where qe is the quantity of water used by a given end-use e (gpd per unit); Mu the mechanical end-use parameter for efficiency class u (e.g., gpm and gallons per flush); Su the fraction of the sectoral end-use within each efficiency class (such as nonconserving, conserving, and ultraconserving); U the intensity of use parameters (e.g., flushes per day per unit, minutes of use per day per unit); K the mechanical parameter representing the rate of leakage; F the fraction of end-uses with leakage; and A the fraction of water-using entities in which end-use is present. Because the structure of end-use demands in a study area at any point in time will not remain constant over the forecast horizon, it is necessary to quantify the effects of various external factors on the parameters of Equation (41)(e.g., Su, U, F, A). For example, future increases in the price of water could

180

Predicting Future Demands for Water

decrease the incidence of leaks (F) in the short run and would also affect the distribution of end-uses among the efficiency classes (Su) in the long run. The other two parameters of the end-use equation (intensity U and presence A) will also be affected by changes in price. Changes in other explanatory variables will also affect the end-use parameters. For example, in the residential sector, in addition to price, the end-use parameters will be affected by variables such as income, household size, housing density, and weather.

5. nonurban; 6. user-specified sector; and 7. total residential. These categories correspond to the housing types used by the US Bureau of the Census. Average rates of water use within each residential subsector are estimated using econometric water-demand models. For the single-family sector, a log–log equation in the form of a multiplicative function of seven independent variables is used:

1.10.4.4.2 Baseline and restricted forecasts Impacts of conservation have to be assessed for different types of conservation measures, including passive, active, and emergency measures as well as future changes in water prices. It is a common forecasting practice to generate one forecast which assumes no active intervention to change future water demands. This forecast is usually referred to as a baseline or unrestricted forecast. Then one or more alternative forecasts are prepared based on assumed active efforts to reduce demands. These forecasts are referred to as restricted forecasts or forecasts with conservation. Some of the forecasting procedures and models described in this and the previous subsections have been incorporated into computer software programs, which structure the input data, provide computational algorithms, and generate forecast outputs. One example of the forecasting program is IWRMAIN.

1.10.4.5 Forecasting Software: The IWR-MAIN Program One of the first known software programs for forecasting urban water use MAIN II (Municipal And Industrial Needs) was developed by Hittman Associates Inc. (1969). The model was based on the residential and commercial water-use research projects carried on at the Johns Hopkins University (Linaweaver et al., 1966; Wolff et al., 1966; Howe and Linaweaver, 1967). During the 1980s, the Institute for Water Resources (IWR) of the US Army Corps of Engineers undertook a substantial research effort to update the model and modify it for easy access on personal computers. The product of this effort was a public-domain software package called IWR-MAIN version 5.1 (Dziegielewski and Boland, 1989) and the IWR-MAIN Water Demand Analysis Software, version 6.0 (Dziegielewski, 1993; Dziegielewski et al., 1996).

1.10.4.5.1 Model structure and procedures The model disaggregates total urban water use by customer sectors, time periods, spatial study areas, and end-use purposes. Through sectoral disaggregation, forecasts of water use can be prepared for major sectors of water users, including residential, nonresidential (constituting manufacturing, commercial, and governmental), public, and other. Demands can be further disaggregated within each sector. Within the residential sector, there are seven subsectors available for forecasting water demands. These include 1. 2. 3. 4.

single-family – 1 attached, 1 detached units; multifamily low density – 2, 3, 4 units per structure; multifamily high density – 5 or more units per structure; mobile homes;

Qr ¼ aIb1 Hb2 Lb3 T b4 Rb5 Pb6 eb7 B

ð41Þ

where Qr is predicted residential water demand in gallons per housing unit per day; I the median household income in $ per year; H the average household size (persons); L the average housing density (units per acre); T the daily-maximum air temperature in farenheit; R the rainfall in inches; P the marginal price of water (including sewer) in $/1000 gallons; B the fixed charge or rate premium (i.e., Nordin’s bill difference) of the water/wastewater tariff in $ per month; a the constant; bi the constant elasticities of explanatory variables; b7 the coefficient of the rate premium (representing the water tariff structure); and e the base of the natural logarithm. The model can use generic water-use equations with default elasticities for explanatory variables in the residential sector which were derived through a meta analysis of empirical literature. Once average water use per household has been estimated by Equation (41), total water demand for a given subsector, season, and year is calculated by multiplying the average use rate Qr by the driver variable (number of households). Nonresidential water use is disaggregated by the model into the following major industry groups: (1) construction; (2) manufacturing; (3) transportation, communications, and utilities; (4) wholesale trade; (5) retail trade; (6) finance, insurance, and real estate; (7) services; and (8) public administration. These eight major industry groups are classified according to the US Department of Commerce Standard Industrial Classification (SIC) codes. Within each major industry group, SIC codes distinguish more homogeneous groups at the twodigit SIC level and even further at the three-digit SIC level. Because no generally applicable demand models exist that contain elasticities for price, labor productivity, cooling degree days, or the other variable for nonresidential water use, IWRMAIN estimates nonresidential water use by multiplying employment with water-use coefficient. Water use per employee coefficients are available for each category and were derived from a sample of about 7000 nonresidential establishments (Table 10). Some estimates in Table 10 represent relatively high values of per employee usage rates (e.g., security and commodity brokers, real estate firms, or legal services), because they represent uses of water other than indoor domestic uses. For example, in large office buildings the component end-uses may include landscape irrigation or makeup water for cooling towers.

Predicting Future Demands for Water Table 10 Water use coefficient for industrial, commercial, and institutional categories of water use Description of SIC categories

Sample size

Construction (SIC 15–17) General building contractors Heavy construction Special trade contractors Manufacturing (SIC 20–39)

246 66 30 150 2790

31 118 20 25 164

252 20 91 62 83 93 174 211 23 116

469 784 26 49 36 2614 37 267 1045 119

10 83 80 395 304 409 182 147 55 226

148 202 178 194 68 95 84 66 36 50

3 32 100 1 10 17 13 31 19 751

68 26 85 5 353 171 40 55 51 43

518 233 1044

46 87 93

Building materials and garden supplies General merchandise stores Food stores Automotive dealers and service stations Apparel and accessory stores Furniture and home furnishing stores Eating and drinking places Miscellaneous retail Finance, insurance, and real estate (SIC 60–67)

56 50 90 198 48 100 341 161 238

35 45 100 49 68 42 156 132 71

Depository institutions Nondepository institutions Security and commodity brokers Insurance carriers Insurance agents, brokers, and service Real estate Holding and other investment offices

77 36 2 9 24 84 5

Food and kindred products Textile mill products Apparel and other textile products Lumber and wood products Furniture and fixtures Paper and allied products Printing and publishing Chemicals and allied products Petroleum and coal products Rubber and miscellaneous plastics products Leather and leather products Stone, clay, and glass products Primary metal industries Fabricated metal products Industrial machinery and equipment Electronic and other electrical equipment Transportation equipment Instruments and related products Miscellaneous manufacturing industries Transportation and public utilities (SIC 40–49) Railroad transportation Local and interurban passenger transit Trucking and warehousing US postal service Water transportation Transportation by air Transportation services Communications Electric, gas, and sanitary services Wholesale trade (SIC50-51) Wholesale trade – durable goods Wholesale trade – nondurable goods Retail trade (SIC 52–59)

Table 10

Continued

Description of SIC categories

Sample size

Services (SIC 70–89)

1878

137

197 300 243 108 42 40 105 353 15 300 55 9

230 462 73 217 69 110 429 91 821 117 106 208

45 5 60 25

212 58 73 106

2 4 6 6 5 2

155 18 87 101 274 112

gped coefficient

62 361 1240 136 89 609 290 (Continued )

181

Hotels and other lodging places Personal services Business services Auto repair, services, and parking Miscellaneous repair services Motion pictures Amusement and recreation services Health services Legal services Educational services Social services Museums, botanical and zoological gardens Membership organizations Engineering and management services Services, NEC Public administration (SIC 91–97) Executive, legislative, and general Justice, public order, and safety Administration of human resources Environmental quality and housing Administration of economic programs National security and International affairs

gped coefficient

gped, gallons per employee per day. Source: Dziegielewski et al. (1996), IWR-MAIN Version 6.0.

1.10.4.5.2 IWR-MAIN conservation forecasts The conservation subroutine of IWR-MAIN 6.0 disaggregates urban water use into 20 end-uses covering residential and nonresidential water uses, both indoor and outdoor. Using Equation (40), each end-use is divided into three classes of efficiency (nonconserving, conserving, and ultraconserving). The rate of use in each of these efficiency classes is defined by the mechanical parameters (M1, M2, and M3). For example, toilets have mechanical parameters of 5.5 gallons per flush (gpf), 3.5 gpf, and 1.6 gpf for nonconserving, conserving, and ultraconserving end-uses, respectively. The percent of sector entities in each efficiency class (S1, S2, and S3) must be determined based on local information. For example, 40% of single-family residential units have toilets with 5.5 gpf, 50% use toilets with 3.5 gpf, and 10% use toilets with 1.6 gpf. The intensity (U) for each fixture defines how frequently a particular end-use occurs on a per household, per employee, or other basis. In the residential sector, an event or flow rate that is defined on a per person basis is multiplied by the average number of persons per household to determine the intensity of the end-use at the household level. For example, a toilet is flushed an average of 5.0 times per person per day. This value, when multiplied by the average number of persons per household in the study area, gives the site-specific intensity value for the toilet end-use. The presence of a particular end-use may vary by forecast year and represents the fraction of units (housing units, employees, or other measures of size) in a water use sector that

182

Predicting Future Demands for Water

have that particular end-use. The ability to adjust the presence factor by forecast year allows the forecaster to account for potential changes in the presence of an end-use. For example, the presence of dishwashers for the single-family sector in the base year may be 76% but could be expected to increase to 84% in 10 years. The conservation savings procedures distinguish between active and passive demand management programs. Active programs include interventions by water providers or other entities. Passive demand reduction is a result of natural shifts toward higher-efficiency classes (e.g., from the standard 3.5 gallon per flush toilet to the ultraconserving 1.6 gallon per flush toilet). The shifts of end-uses toward higher classes of efficiency are brought about primarily by plumbing codes that require increased efficiency in water-using fixtures which affect new construction, remodeling, and customer-initiated retrofitting. Shifts from less efficient to more efficient pools of end-uses are also expected to occur naturally over time as technology continually improves. The natural shifts toward higher efficiency pools (often accelerated by plumbing or efficiency codes) are determined using appropriate rates of movement. These movements predict the form in which these shifts take place and additionally show the rate at which they occur. Conservation measures may be defined as long-term, shortterm, or emergency (restricted use) programs. Long-term conservation programs permanently shift the number of entities into higher efficiency classes. These include all plumbing codes (passive programs) as well as active programs. An example of an active program that would be considered a longterm conservation program is a retrofit campaign in which low-flow showerheads are installed. Showerheads usually last about 15 years, after which it is assumed that they will be replaced with units that are at least as efficient, if not more efficient. Short-term conservation measures are implemented during periods of water shortages. These measures target the end-use fixtures by temporarily shifting them into a higher efficiency class. An example would be the distribution of toilet dams. These devices would temporarily shift users into a higher efficiency level but are unlikely to remain in the toilets for a long period of time. Although the short-term conservation measures focus on increasing the efficiency of end-uses, the emergency or restricted use measures focus on temporarily altering the behavior of water-using entities. Like the short-term programs, the restricted use programs evaluate potential reductions in demand which could be enacted during periods of water shortages. A restricted use conservation program might entail, for example, restricting the days when the residents of a community can wash their cars. Restrictive programs target specific end-uses and achieve water savings by using alternate (restrictive) values for presence factors, leakage percents, and intensities.

1.10.5 Example of a Regional Multisector Forecast This example taken from Dziegielewski and Chowdhury (2008) illustrates the development of future water-demand

scenarios for geographical areas that encompass the 11-county regional planning area of Northeastern Illinois, including the counties of Boone, Cook, DeKalb, DuPage, Kane, Kankakee, Kendall, Grundy, Lake, McHenry, and Will (Figure 2). In 2005, total resident population of the area was estimated at 8 743 900 persons and a total of 4 355 200 persons were employed in the local economy. Nearly 96% of the population was served by about 530 public water-supply systems and nearly 400 000 residents relied on private wells. Other water users included 12 large power plants, 352 golf courses, and about 30 000 acres of irrigated cropland. The study generated three water-demand scenarios by major user sectors and geographical service areas within the region that were extended to the year 2050. The scenarios were formulated to represent growth assumptions under current trends (CTs or baseline scenario) as well as under less resource-intensive (LRI or low-growth scenario) and more resource-intensive (MRI or high-growth scenario).

1.10.5.1 Water-Use Relationships The historical data on water withdrawals were used to estimate water-demand relationships that expressed water demand as a function of relevant explanatory variables. Table 11 lists the demand drivers and estimated elasticities of the explanatory variables for each demand sector. Because the dependent and, in most cases, independent variables were converted to natural logarithms, the coefficients represent constant elasticities of water demand. Accordingly, the elasticity of marginal price in the public-supply sector was estimated to be  0.1458, indicating that a 1.0% increase in price is expected to result in a 0.145 8% decrease in demand. The three forecast scenarios were defined by different sets of assumed conditions regarding the future values of demand drivers and explanatory variables. Table 12 compares several key assumptions that were used in constructing the three scenarios. Table 13 provides a summary of the future scenarios of average day water withdrawals for six categories of users within the four major sectors. For 2005, both the reported values and weather-adjusted values are shown. The lower panel of Table 13 shows the sum of total withdrawals with and without the once-through flow for power generation and gross per capita withdrawals without once-through flows. This distinction is made because the very high volumes of water withdrawals for once-through cooling are not directly comparable to withdrawals by other sectors. The three future water-demand scenarios show that total water withdrawals in the 11-county area of Northeastern Illinois will continue to increase to meet the demands of growing population and the concomitant growth in the economy of the region. However, the growth in total water demand could be faster or slower depending on which assumptions and expectations about the future conditions will prevail. Under the baseline (CT) scenario, by 2050, total water withdrawals (excluding water withdrawn for once-through cooling in electric power plants) would increase above the 2005 level by 35.8%, or 530 mgd. During the same period of time, total population is projected to increase by nearly 3 370 000, or 38.5%. This implies that water demand would grow slightly

Predicting Future Demands for Water

183

W I S C O N S I N BOONE

MCHENRY

LAKE

I

L

L

Chicago

DUPAGE

KANE

I

N

O

I

S

e

r

DEKALB

Lake Michigan

Fox

Riv

COOK KENDALL

WILL Ka nk a

N

*Groundwater is also used within these areas in some cases.

k Ri

GRUNDY

ee

10 MILES

INDIANA

ve r

KANKAKEE

Source: Chicago Metropolitan Agency for Planning

Figure 2 Regional forecast area in Northeastern Illinois.

slower than the region’s population. Gross per capita water withdrawals (i.e., total withdrawals by all sectors divided by total population) during the dry year of 2005 were estimated at 182.8 gpcd. Under normal weather conditions, the 2005 demands would have been 169.3 gpcd. In the baseline (CT) scenario, the gross per capita usage would decrease to 166.0 gpcd by 2050. This relatively unchanged per capita rate is a result of assumptions about gradual increases in water prices and a continuation of the historical trend in water conservation. Under the key assumptions of the high-growth (MRI) scenario, future water demands would grow faster than total population if income grows at a somewhat higher rate than under the baseline scenario, if future prices of water do not grow faster than inflation, if no additional gains in water conservation are achieved, and if more population growth takes place in the collar counties of Kane, Kendall, and McHenry in single-family housing. Under these conditions, by 2050, total water withdrawals would increase above the 2005 level by 64.1%, or 949.1 mgd. The growth of water demand would exceed the rate of population growth because of the increasing gross per capita usage rate. By 2050, it would increase to about 200.6 gpcd, as compared to the 2005

weather-normalized rate of 169.3 gpcd. In a sense, the high-growth scenario could be viewed as a warning that there is a possibility of a large increase of water demands in the future. Under the key assumptions of the low-growth (LRI) scenario, future demands would grow significantly slower than population if income grows at a somewhat slower rate than under the baseline scenario, if future prices of water grow significantly faster than inflation, if additional gains in water conservation are achieved, and if more population growth takes place in the urbanized counties of Cook and DuPage in multifamily housing. Under these conditions, by 2050, total water withdrawals would increase above the 2005 level by about 7.2%, or 107.2 mgd. The growth of water demand would be much slower than the rate of population growth because of the decreasing gross per capita usage rate. By 2050, it would decrease to about 131.1 gpcd, as compared to the 2005 weather-normalized rate of 169.3 gpcd. This scenario could be interpreted as a future outcome which would require an intervention in order to maintain a slower growth of demand. This intervention would likely require monitoring and management of water demand and making investments in long-term efficiency of water use.

184

Predicting Future Demands for Water

Table 11

Drivers of water demand and estimated elasticities of explanatory variables

Demand sector

Demand driver

Explanatory variables

Elasticity/coefficient

Public supply

Population served

Power generation

Gross electric generation

Air temperature Precipitation (growing season) Employment fraction Marginal price of water Median household income Conservation trend Unit-use coefficients

Industrial and commercial

Employment

Agricultural and irrigation

Irrigated acres Livestock counts Population

1.0951 –0.0949 0.0931a –0.1458 0.2845 –0.0593 0.67–0.89b 10.8–78.9c 0.3298 –0.0896 0.0279a –0.1077a 0.0032a –0.0074a 1.000 0.03–35.0d 1.6238 –0.2186 0.3499 –0.0325

Domestic self-supplied

Cooling degree days Precipitation (growing season) Manufacturing employment (%) Transportation employment (%) Fraction of self-supplied (%) Conservation trend (linear) Rainfall deficit Unit-use coefficients Air temperature Precipitation (growing season) Median household income Conservation trend

a

Dependent variables are in linear form. All other coefficient represent constant elasticities. The values represent unit withdrawal coefficients in gallons per kilowatt-hour of gross generation in plants with closed-loop cooling systems. c The values represent unit withdrawal coefficients in plants with open-loop once through cooling systems. d The values represent unit use coefficient per animal type. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008. b

Table 12

Scenario assumptions for factors affecting future water demands in the 11-county area of Northeastern Illinois

Factor

CT (baseline)

LRI scenario

MRI scenario

Distribution of population of growth Mix of commercial/industrial activities Median household income Demand for electricity

Official projections

Shift to Cook and DuPage counties Decrease in water-intensive activities Growth of 0.7% yr1 9.61 MWh per capita per year þ no growth No new power plants, 3 units retired, 2 plants convert to closed-loop cooling 50% higher rate than historical trend Growth of 2.5% yr1 Decreasing cropland þ no new golf courses

Shift to Kane, Kendall, and McHenry counties Increase in water-intensive activities Growth of 1.0% yr1 9.61 MWh per capita per year þ 0.56% per year growth Two new power plants in study area with closed-loop cooling

Power generation

Current trends Growth of 0.5% yr1 9.61 MWh per capita per year þ 0.56% per year growth No new plants within study area, 3 units retired

Water conservation

Historical trend

Future water prices Irrigated land

Growth of 0.9% yr1 Constant cropland þ new golf courses: 10 per decade

No extension of historical trend Constant in 2005$ Constant cropland þ new golf courses: 20 per decade

CT, current trends; LRI, less resource-intensive; MRI, more resource-intensive. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

1.10.5.2 Effects of Key Forecast Assumptions The plausibility of the scenario forecasts depends on how reasonable the assumptions about future changes in the explanatory variables and the size of their effects on water demand are. In order to verify the forecasts of total or sectoral water use in the future, it is helpful to re-analyze the resultant

forecasts in terms of the underlying effects of demand drivers and the unit usage rates. The total change in demand is the result of the change in the value of the driver (e.g., population served, employment, and irrigated acreage) and the change in unit usage rate. Table 14 compares the resultant unit usage rates for the major demand sectors for each scenario. The changes in future values of these rates are a result of changes in

Predicting Future Demands for Water Table 13

185

Summary of water withdrawal scenarios for Northeastern Illinois (in mgd)

Sector

2005-R

2005-Na

2050-CT

2050-LRI

2050-MRI

Public supply Self-supplied I&C Self-supplied domestic Irrigation and agriculture Power plants (makeup) Power plants (through flow) Total – all sectors Total w/o through-flow power

1255.7 191.6 36.8 62.0 52.3 4207.2 5805.6 1598.4

1189.2 162.4 31.8 44.6 52.3 4207.2 5687.5 1480.3

1570.2 291.6 41.2 55.4 52.3 3830.2 5840.9 2010.7

1217.9 222.1 37.3 43.8 66.4 2472.3 4059.8 1587.5

1837.2 391.4 49.3 60.7 90.8 3830.2 6259.6 2429.4

a

For comparison with future values, the 2005 withdrawals were adjusted by the model to represent normal weather conditions. R, reported; N, normal weather; CT, current trends; LRI, less resource-intensive; MRI, more resource-intensive; mgd, gallons per capita per day; 1.0 mgd, 3784.4 m3. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

Table 14

Summary of unit usage rates by sector

User sector

2005-R

2005-Na

2050-CT

2050-LRI

2050-MRI

Public supply (gpcd) Self-supplied I&C (gped) Self-supplied domestic (gpcd) Irrigation and agriculture (in yr1) Power plants (makeup)(gal. kWh1) Power plants (through-flow)(gal. kWh1) Gross per capita rates (gpcd)

150.4 130.6 93.6 15.3 0.86 41.4 182.8

142.1 109.3 81.1 11.0 0.86 41.4 169.3

134.9 120.1 86.4 11.5 0.86 40.2 166.0

104.7 90.6 78.3 9.1 0.81 33.3 131.1

157.9 155.2 103.5 12.5 0.77 40.2 200.6

a

For comparison with future values, the 2005 withdrawals were adjusted by the model to represent normal weather conditions. gpcd, gallons per capita per day; gped, gallons per employee per day; in yr1, irrigation water application depth in inches per year; gal. kWh1, water use for thermoelectric cooling in gallons per kilowatt-hour. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

Table 15

Effects of driver and unit rate changes on future water demand (CT scenario)

Sector

Public supply Self-supplied I&C Self-supplied domestic Irrigation and agriculture Power plants (makeup) Power plants (through flow) Total (w/o through flow)

2005 demand (mgd)

1189.2 162.4 31.8 44.6 52.3 4207.2 1480.3

2005–2050 increase in demand

Effect of change in unit rates

Effect of change in driver count

mgd

%

mgd

%

mgd

%

381.0 129.2 9.4 10.8 0.0 –377.0 530.4

32.0 79.5 29.6 24.2 0.0 –9.0 35.8

–83.5 37.5 2.6 2.2 0.0 –109.8 –41.1

–7.0 23.1 8.2 4.9 0.0 –2.6 –2.8

464.5 91.7 6.8 8.6 0.0 –267.2 571.5

39.0 56.4 21.4 19.3 0.0 –6.4 38.6

CT, current trends; mgd, million gallons per day. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

key explanatory variables such as income, price, or conservation trend. The effects of weather conditions are shown by comparing the observed and weather-normalized usage rates for 2005. Only normal-weather predictions are shown for the future years. Table 15 compares the effects of growth (or change) in demand drivers, which are separated from the effects of

projected changes in unit usage rates. The 2005–50 increments in demand are decomposed into the effect of growth (or change) of demand driver and the effect of projected changes in unit usage rates. For example, for the public-supply sector, the 32.0% increase in demand is a net result of a 39% increase in population served and a 7% decrease in per capita rate of water use.

186 Table 16

Predicting Future Demands for Water Effects of key assumptions on per capita rates in public supply sector

Forecast assumption

2005-Na

Public supply (gpcd) Total 2005–2050 effect Effects of individual variables

142.1

Retail price of water Median household income Conservation trend Employment/population ratio Differential growth

2050-CT

2050-LRI

2050-MRI

134.9 –7.2

104.7 –37.4

157.9 þ 15.8

 8.2 þ 11.5 –11.2 þ 1.2 –0.5

–18.4 þ 6.5 –23.3 þ 1.0 –3.2

0.0 þ 18.1 0.0 þ 1.4  2.7

a

For comparison with future values, the 2005 withdrawals were adjusted by the model to represent normal weather conditions. CT, current trends; LRI, less resource-intensive; MRI, more resource-intensive; N, normal weather; gpcd, gallons per capita per day. From Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planning, Chicago, IL, USA, 15 June 2008.

The change in unit usage rates is a result of changes in the future values of explanatory variables. Table 16 shows the effects of future changes in the values of four variables on per capita rates in public-supply sector. For the low-growth scenario, the per capita rate of usage is projected to decrease by 37.4 gpcd. This value represents a net effect of the increase in marginal price, increase in median household income, effect of conservation, and effect of the changing ratio of employment to resident population. One additional component of the change in gpcd is the effect of differential growth in population across the 37 geographical areas used in the forecast. For the baseline scenario, this effect is  0.5 gpcd; it would be zero if population served in all areas was projected to grow at the same rate. Making the effects of these different factors explicit allows the stakeholders who participate in the planning process to assess the reasonableness of the forecast and to decide which scenarios should be taken into account when formulating water-supply plans.

1.10.5.3 Effects of Future Climate Climate models indicate that by 2050 in Illinois there may be an average annual normal temperature departure of up to þ 6 F, and a possible departure from normal annual precipitation in a range of  5 to þ 5 in yr1 from the 1971–2000 long-term normal values (ISWS, 2007). Due to the nature of climate scenarios, no probabilities can be placed on the possible ranges of future air temperature and precipitation. The changes in normal annual temperature and precipitation would also result in average-weather changes during the growing season. The temperature increase of 6 F will also apply to the summer growing season. The distribution of precipitation changes is expected to range from þ 2.5 to  3.5 in during the growing season. The effects of these changes on future demands will vary by user sector, depending on each sector’s sensitivity of water withdrawals to air temperature and precipitation. The results show that future demands in all sectors are likely to be higher if future annual average air temperature increases and/or annual precipitation decreases. If, by 2050, temperature increases by 6 F, total withdrawals would increase

by 178.0 mgd (9.1%) above the baseline (CT) scenario values. The largest increase in total withdrawals above the baseline scenario would be 229.5 mgd (or 11.7%) by 2050, resulting from the combined effect of the temperature increase and a decrease in summer precipitation.

1.10.5.4 Forecast Summary The scenario forecasts of future water demands in the 11county planning area in Northeastern Illinois revealed the possibility of potentially large increases in total water withdrawals by 2050. The baseline (CT) and high-growth (MRI) scenarios when viewed in the context of regional supply limitations make a compelling case for the need to manage regional water demands. Total withdrawals (excluding oncethrough flows in power plants) in 2050 under baseline scenario would increase by 530.4 mgd and the increase could reach as high as 949.1 mgd under high-growth scenario. Meeting these additional demands would require large capital outlays for water infrastructure and would likely have significant impacts on some of the regional sources of water supply, especially groundwater aquifers and local rivers, which would create increased demands on water from Lake Michigan.

1.10.6 Conclusion The prospect of climate change, which could alter meteorological and hydrological regimes and possibly result in lower water availability and higher demands, is challenging water managers to search for effective ways to satisfy future demands without jeopardizing the long-term sustainability of current water resources systems. Credible long-term forecasts of water demand can help water planners to achieve an efficient allocation of future water supplies among competing uses. When predicting future demand for water, it is important to recognize that water is an economic good and, therefore, its future use will be responsive to changes in future economic conditions. The economic considerations should help water planners to achieve efficient and equitable use of water supplies while encouraging conservation and protection of water resources.

Predicting Future Demands for Water

This chapter described data and methods for developing a clear and credible forecast of future water demand which would enhance its acceptability by decision makers. The main characteristics of such a forecast include a high level of disaggregation of demand, use of econometric water-use models which conform to the economic theories of production and consumption, and provision of explicit and plausible forecasting assumptions. A credible forecast cannot be developed without first examining the historical data on water use. Obtaining valid data on water use for a sufficiently long period of time (usually 15–20 years) is usually the most time-consuming task of a forecasting project. The data must allow for disaggregation of total water use into different categories of demand and also have appropriate geographical and temporal resolution to permit the development of water-use relationships which would capture the effects of variables which affect water demand and operate at different scales of space and time. Additional large effort is required in obtaining data on explanatory variables. Once appropriate data are assembled, the next challenge is the development of water-use relationships, usually by applying statistical methods such as multiple regression. The criteria for deriving a empirical model which is useful for forecasting are somewhat different than those in a typical econometric studies where researcher wishes to know which variables are statistically significant and if the resultant model conforms to economic theory. A useful forecasting model requires not only an appropriate model specification but also accurate estimates of the regression coefficients (or elasticities) for the explanatory variables. The forecaster must develop fairly strong expectations about the size of the regression coefficients for key explanatory variables such as price, income, air temperature, and precipitation. Expectations about the sign (positive or negative) of the explanatory variables come directly from economic theory and the underlying physical relationships. The next step in developing a credible forecast involves the development of forecasting assumptions which are the basis and the main component of the forecast. The most critical are the assumptions about the future values of explanatory variables. Where possible, forecasts of water demand should be based on official or consensus forecast of population and economic growth in the study area. The future values of other explanatory variables (e.g., price of water, income, or climate) should be derived based on explicit and plausible assumptions. The uncertainty caused by forecasting assumptions can be addressed by using forecast scenario, sensitivity analysis, or developing probability forecasts. Finally, once the forecasted quantities of future water demand are prepared, the forecaster should provide a postforecast analysis of results. In essence, such an analysis should identify and quantify the effects and relative contributions of individual forecasting assumptions on the final results. For example, the change in water demand during the forecast period can be separated into the effects of population growth, changes in the price of water, income, climate, and other explanatory variables. The post-forecast analysis can help ensure that the forecast is understood and accepted by decision makers.

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References Avery C (1999) Estimated water withdrawals and use in Illinois, 1992. Open-File Report 99–97. Urbana, IL: US Geological Survey. Backus CE and Brown ML (1975) Water requirements for solar energy. In: Gloyna EF, Woodson HH, and Howard Drew R (eds.) Water Management for the Electric Power Industry, Water Resources Symposium #8, pp. 270–279. Austin, TX: Center for Research in Water Resources, The University of Texas at Austin. Baumann D, Boland JJ, and Hanemann WM (1998) Urban Water Demand Management and Planning. New York: McGraw-Hill. Baumann DD, Boland JJ, and Sims JH (1984) Water conservation: The struggle over definition. Water Resources Research 20(4): 428--434. Boland J (1998) Forecasting urban water use: Theory and principles. In: Baumann DD, Boland JJ, and Hanemann M (eds.) Urban Water Demand Management and Planning, ch. 3, pp. 77–94. New York: McGraw-Hill. Boland JJ (1985) Forecasting water use: A tutorial. In: Torno HC (ed.) Computer Applications in Water Resources, pp. 907--916. New York, NY: The Society. Boland JJ, Dziegielewski B, Baumann DD, and Opitz EM (1984) Influence of price and rate structures on municipal and industrial water use. IWR Report 84-C-2. Fort Belvoir, VA: US Army Engineer Institute for Water Resources. Boland JJ, Dziegielewski B, Baumann DD, and Turner C (1982) Analytical bibliography for water supply and conservation techniques. IWR Report 82-C-07. Fort Belvoir, VA: US Army Engineer Institute for Water Resources. Blake NM (1956) Water for the Cities: A History of the Urban Water Supply Problem in the United States. Syracuse, NY: Syracuse University Press. Breusch TS and Pagan AR (1979) A simple test for Heteroskedasticity and random coefficient variation. Econometrica 47: 1287--1294. Croley TE II, Patel VC, and Cheng MS (1975) The Water and Total Optimization of Wet and Dry–Wet Cooling Towers for Electric Power Plants. Iowa City, IA: Iowa Institute of Hydraulic Research, University of Iowa. Deaton A and Muellbauer J (1999) Economics and Consumer Behavior. Cambridge: Cambridge University Press. De Rooy J (1974) Price responsiveness of the industrial demand for water. Water Resources Research 10(3): 403--406. Dupond D and Renzetti S (1998) Water use in the Canadian food processing industry. Canadian Journal of Agricultural Economics 46: 1--10. Dziegielewski B (1993) IWR-MAIN 6.0: A tool for demand management and planning. Journal of American Water Works Association (Aqualink Department) 85(8): 24. Dziegielewski B and Baumann DD (1992) Benefits of managing urban water demands. Environment 34(9): 6--11. 35–41. Dziegielewski B, Baumann DD, and Boland JJ (1983) Prototypal application of a drought management optimization procedure to an urban water supply system. IWR Report 83-C-4. Fort Belvoir, VA: US Army Engineer Institute for Water Resources. http://stinet.dtic.mil/oai/oai?&verb=getRecord&metadataPrefix=html &identifier=ADA138473 (accessed April 2010). Dziegielewski B and Bik T (2006) Water use benchmarks for thermoelectric power generation. Prepared for the United States Geological Survey. Southern Illinois University at Carbondale, 15 August 2006. Dziegielewski B and Boland JJ (1989) Forecasting urban water use: The IWR-MAIN model. Water Resources Bulletin 25(1): 101--109. Dziegielewski B and Chowdhury FJ (2008) Regional water demand scenarios for Northeastern Illinois: 2005–2050. Project Completion Report. Prepared for the Chicago Metropolitan Agency for Planing, Chicago, IL, USA, 15 June 2008. Dziegielewski B and Kiefer JC (2006) U.S. water demand, supply and allocation: Trends and outlook. IWR Report 2007-R-3. A white paper prepared for the U.S. Army Corps of Engineers Institute for Water Resources, Alexandria, VA, USA, 22 December 2006. Dziegielewski B, Opitz EM, Kiefer JC, and Baumann DD (1993) Evaluation of Urban Water Conservation Programs: A Procedures Manual, xxi þ 274pp., Book No. 0-89867-676-2. Prepared for California Urban Water Agencies, Sacramento, CA. Denver, CO: American Water Works Association. Dziegielewski B, Opitz EM et al. (1996) IWR-MAIN Water Demand Analysis Software. Users Manual and System Description. Carbondale, IL: Planning and Management Consultants. Dziegielewski B, Rodrigo D, and Opitz E (1990) Commercial and Industrial Water Use in Southern California. Carbondale, IL: Planning and Management Consultants. Dziegielewski B, Sharma SC, and Margono H (2005) Models for long-term forecasts of public supply water use. In: Proceedings of the XIIth IWRA World Congress on Water Resources. New Delhi, India, 22–25 November 2005. New Delhi, India: Central Board of Irrigation and Power.

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Dziegielewski, B, Strus CA, and Hinckley RC (1993) End-use approach to estimating water conservation savings. In: CONSERV93: The New Water Agenda. American Water Works Association, Denver, Colorado. Grigg NS (1986) Urban Water Infrastructure: Planning, Management, and Operations. New York: Wiley. (original from the University of Michigan, MI). Epsey M, Epsey J, and Shaw WD (1997) Price elasticity of residential demand for water. Water Resources Research 33(6): 1369--1374. Ferguson B and Whitney A (1993) Demand reduction in response to drought: The city of Santa Barbara experience. In: CONSERV93 Proceedings, ASCE, AWWA, AWRA, pp. 429–437. Foster HS Jr and Beattie BR (1979) Urban residential demand for water in the United States. Land Economics 55(1): 43--58. Gardiner V and Herrington P (1986) Water Demand Forecasting: Proceedings of a Workshop. Exeter: Short Run Press. Glejser H (1969) A new test for heteroscedasticity. Journal of the American Statistical Association 64: 316--323. Hanemann WM (1998) Determinants of urban water use. In: Baumann DD, Boland JJ, and Hanemann M (eds.) Urban Water Demand Management and Planning, ch. 2, pp. 31–75. New York: McGraw-Hill. Hittman Associates, Inc. (1969) Forecasting Municipal Water Requirements, Volume 1: The MAIN II System, PB 190275. Columbia, MD: Hittman Associates, Inc. Howe CW (1982) The impact of price on residential water demand: Some new insights. Water Resources Research 18(4): 713--716. Howe CW and Linaweaver FP Jr (1967) The impact of price on residential water demand and its relation to system design and price structure. Water Resources Research 3(1): 13--32. Hutson SS, Barber NL, Kenny JF, Linsey KS, Lumia DS, and Maupin MA (2004) Estimated use of water in the United States in 2000. US Geological Survey. USGS Circular 1268. http://water.usgs.gov/pubs/circ/2004/circ1268/ (accessed April 2010). ICWE (International Conference on Water and the Environment) (1992) The Dublin statement on water and sustainable development. http://www.un-documents.net/ h2o-dub.htm (accessed April 2010). IPCC (2008) Definition of terms used with the DDC pages. http://www.ipcc-data.org/ ddc_definitions.html (accessed July 2010). ISWS (Illinois State Water Survey) (2007) Tomorrow’s climate – future scenarios: CO2 concentrations. http://www.sws.uiuc.edu/wsp/climate/ ClimateTom_scenarios_co2.asp (accessed April 2010).

IWR (Institute for Water Resources) (2001) An evaluation of the risk of water shortages in the Lower Peninsula, Virginia. Prepared by Werick WJ, Boland JJ, Gilbert J, Dziegielewski B, Kiefer J, Massmann J, and Palmer RN. IWR Special Report. Alexandria, VA: US Army Corps of Engineers. Linaweaver FP, Jr., Beebe JC, and Skrivan FA (1966) Data Report of the Residential Water Use Research Project, Johns Hopkins University, Department of Environmental Engineering Science, Baltimore, MD. Metcalf L (1926) Effects of water rates and growth in population upon per capita consumption. Journal of the American Water Works Association 15(1): 1--20. Nieswiadomy ML (1992) Estimating urban residential water demand: Effects of price structure, conservation, and education. Water Resources Research 28(3): 609--615. Platt RH (1993) Water demand management. Commentary. Environment 35(3): 2--3. Ramsey JB (1969) Tests for specification errors in classical linear least-squares regression analysis. Journal of the Royal Statistical Society Series B 31: 350--371. Renzetti S (1988) An econometric study of industrial water demands in British Columbia, Canada. Water Resources Research 24(10): 1569--1575. Renzetti S (1992) Estimating the structure of industrial water demands: The case of Canadian manufacturing. Land Economics 69(2): 181--188. Renzetti S (1993) Examining the difference in self- and publicly supplied firms’ water demands. Land Economics 69(2): 181--188. Renzetti S (2002) The Economics of Water Demands. Norwell, MA: Kluwer. Solley WB, Pierce RR, and Perlman HA (1998) Estimated Use of Water in the United States in 1995, U.S. Geological Survey Circular 1200. Denver, CO: US Geological Survey. UNCED (United Nations Conference on Environment and Development) (1992) Agenda 21, Protection of the Quality and Supply of Freshwater Resources: Application of Integrated Approaches to the Development, Management and Use of Water Resources, ch. 18. Brazil: Rio de Janeiro. USDA (United States Department of Agriculture) (2009) 2007 Census of Agriculture, Summary and State Data Volume 1, Geographic Area Series Part 51. Washington, DC: National Agricultural Statistics Service. White H (1980) A heteroskedasticity-consistent covariance matrix estimator and a direct test for heteroskedasticity. Econometrica 48: 817--838. WHPA and Dziegielewski B (2008) Water Demand Scenarios for the East-Central Illinois Planning Region. Bloomington, IN: Wittman Hydro Planning Associates. Wolff JB, Linaweaver FP, Jr., and Geyer JC (1966) Water Use in Selected Commercial and Institutional Establishments in the Baltimore Metropolitan Area. Johns Hopkins University.

1.11 Risk Assessment, Risk Management, and Communication: Methods for Climate Variability and Change C Brown, University of Massachusetts, Amherst, MA, USA KM Baroang, International Research Institute for Climate and Society, Palisades, NY, USA & 2011 Elsevier B.V. All rights reserved.

1.11.1 1.11.2 1.11.2.1 1.11.2.2 1.11.2.3 1.11.3 1.11.4 1.11.4.1 1.11.4.2 1.11.4.2.1 1.11.4.2.2 1.11.4.2.3 1.11.5 References

Introduction Background on Risk Assessment and Management Hazard Characterization Risk Assessment Risk Management Risk Management versus Consequence Management: The Upside of Risk Climate Risk Risk and Nonstationarity: Uncertain Information and Unreliable Probability Estimates Climate Consequence Management with Decision Scaling: An Approach Designed for Uncertain Information Step 1: Vulnerability and uncertainty identification Step 2: Consequence assessment Step 3: Consequence and uncertainty management Conclusion

1.11.1 Introduction The subject of risk is a familiar one for the water resources engineer and the water planner. When a harmful event is possible but not certain, there is risk. In the world of water resources harmful events are largely, but not solely, associated with precipitation: too much, too little, and too variable. Other aspects of the hydrologic cycle modulate this basic signal as do various human activities, but we largely depend on precipitation and it is uncertain. Risk is inherent to the profession. Yet there does not seem to be a predominant method for addressing risk in the profession. This fact gains prominence due to climate change, nonstationarity of the hydrologic record, and the potential for changing and growing risks in the future. In the field of water resources, research has largely proceeded from a scenario-based rather than a riskbased framework of analysis. The primary direction of analysis has been simulation of a few possible future scenarios and assessment of impacts corresponding to these scenarios. Unfortunately, the uncertainty of these scenarios in comparison to other possible futures makes this approach unhelpful for risk assessment. It may be that the approach to risk is so institutionalized that all remains is discourse on the probabilities of extreme events and how to produce better estimates of them. The cost of this institutionalization is an accompanying ossification of the way in which risk is addressed. Climate change implies that our tried and true methods to reduce uncertainties to produce estimates of risk may no longer serve us well. For this reason, it is an opportune moment to review the treatment of risk in water resources, from assessment to management and finally to communication. Interestingly, the rise in importance of risk management as a framework to plan for a changing climate is accompanied by a questioning as to

189 190 190 191 191 193 193 194 195 195 196 197 198 199

whether the prevailing approaches to risk management are still valid. As we shall see, standard approaches to risk assessment and risk management as employed in the water sector utilize estimates of probability density functions (PDFs) associated with hydrologic variables. There is reasonable concern as to whether we have the ability to assign such probabilities with any confidence given to the small sample size associated with extreme events, which are the events of interest. There is further concern as to whether use of the historic record to estimate hydrologic design variables is appropriate for designs prepared for the future climate. In order to explore these issues, we require an understanding of what risk assessment and management mean and then explore why climate change may necessitate a refinement of this approach. In this chapter, we will review basic risk concepts, some history of its development in water resources engineering and planning, visit some emerging approaches to risk associated with deep uncertainties such as those associated with climate change, and finally propose a process for addressing climate risks to water resources systems. In doing so, we find that a process for climate risk is a special application for the general risk assessment and management framework as presented here. Because the terminology associated with risk is used to mean different things in different communities, it is useful to begin with some basic definitions. The definitions listed here are generally consistent with usage in water resources and hydrology and to some extent with the natural hazards community (Plate, 2002; UNDRO Office of the United Nations Disaster Relief Coordinator, 1991): Hazard. The probability of an event that causes failure or a negative effect on a community or system. Risk. The product of the probability of a hazardous event occurring and the impact or consequence of that event; risk

189

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can increase if either the probability increases or the consequences of a hazard become more severe. Vulnerability. The characteristics of a community or system that cause them to be susceptible to adverse outcomes when exposed to a particular event. Resilience. The capacity of a community or system to recover from an adverse outcome due to a hazard and return to an acceptable level of functioning. Careful consideration of the definition of risk raises a topic of interest that is often not addressed in water resources. This is the idea that there are opportunities presented by uncertain events that have positive consequences. Our typical focus is on the tail of the event probability distribution that is associated with negative consequences, such as droughts or floods. However, there can also be beneficial opportunities associated with the tails of probability distributions. Consider the case of a farmer engaged in rain-fed agriculture planning the crop pattern for the season ahead. If that farmer plans a pattern optimal for the average seasonal rainfall total but not ideal for conditions that are below normal, then there is a risk associated with below-normal rainfall. If the rainfall is below normal then the crop yield will be reduced. The farmer could reduce his risk exposure to drought by choosing a crop pattern that is well suited for below-normal rainfall. However, the farmer would now forgo the opportunities associated with above-average rainfall, say a bumper crop of a water-intensive, high yielding cash crop. The relationship between risk and opportunity will be discussed later in this chapter.

yielding events, when they occur. Note that the consequences of an event are contingent on the risk management steps taken or not taken, and so C(x) can be specified as C(x|D), where D represents a decision. For example, the consequences associated with a flood of stage x is a function of the decision, D, to build a levee of some level or none at all. In the typical approach, the estimation of risk then requires the estimation of two functions, the probability of occurrence of the events of interest, f(x), and the estimation of the consequences of that event for a given decision, C(x|D). The identification of hazards and the estimation of f(x) associated with those hazards is the first step of addressing risks, which is called hazard characterization. This process is discussed in more detail below. While f(x) is typically thought of as acts of nature, in the case of hydrology human action can influence the characteristics of stream flow. For example, significant development of natural land surfaces to conditions of partially or completely impervious surface would likely change the frequency of high flow events. Decisions to preserve land surfaces in natural states or to develop them with low-impactdevelopment designs could similarly change the frequency of high flow events. Thus, one could consider a PDF of stream flow that is conditional on decisions, such as land use, that are made that affect stream flow characteristics, f(x|d). The next step is risk assessment which consists of the estimation of the expected value of the consequences, C(x|D). This is accomplished by integrating over the PDF of x:

 ¼ RðxÞ

Z

N

CðxjDÞf ðxÞdx

ð1Þ

0

1.11.2 Background on Risk Assessment and Management There are a variety of processes with different steps that have been proposed for risk assessment and management. Some place risk assessment as a single step within a larger risk management framework. Here, we describe the process of three general steps which cover the various processes that are consistent with the themes in most approaches. We also highlight particular topics not typically covered in depth, namely opportunities, residual risk, and surprise. The three general steps can be described as a hazard characterization step, a risk assessment step, and a risk management step. The mathematical description presented here draws from previous formalizations of risk, for example, Plate (2004) and Lund (2002). Risk consists of impacts or consequences that result from an event that may occur with some probability. It can be summarized in the following equation:

RðxÞ ¼

Z

where R is the expected risk for a given decision, D. The process of risk assessment produces a risk associated with a hazard or range of hazards. In the final step of risk management, one can then attempt to make a decision or decisions that reduce the unwanted risks. One can then evaluate the expected risks for the set of possible decisions, Di, i ¼ 1,y, N, for N possible decisions and select the risk-minimizing decision:

min Z ¼

Z

D

N

CðxjDÞf ðxÞdx

ð2Þ

0

Considering decisions related to changes to stream flow characteristics and thus the PDF of the stream flow, the equation can be more generally specified as

min Z ¼ D;d

Z

N

CðxjDÞf ðxjdÞdx

ð3Þ

0

N

CðxÞf ðxÞdx

0

where f(x) is the PDF of the event and C(x) is the consequences associated with that event on the system of interest. For the hydrologic risks, f(x) is the PDF of a hydrologic event such as a flood of the stage (height), x. The consequence function yields the costs or damages associated with a flood of stage x. We use the term consequence function in place of the more common damage function to highlight the potential for benefits that may arise from opportunities, uncertain benefit

This is one possible mathematical specification of the quantification of risks. One could consider this as an optimization, either formally with an optimization model that can be used to evaluate risk management alternatives or informally through a decision process to achieve a similar objective. The actual processes involved in hazard characterization, risk assessment, and risk management are more difficult. While there is much focus within the research community on the description of f(x) in terms of the selection of the distribution type and methods employed for the fitting of

Risk Assessment, Risk Management, and Communication: Methods for Climate Variability and Change

parameters, the importance of correctly characterizing CðxjDÞ and of developing a wide range of options for managing risks, Di, are perhaps of greater importance given to our growing awareness of the limitation in correctly specifying f(x).

1.11.2.1 Hazard Characterization The first step is one of identifying and characterizing hazards. This is an exploratory phase in which the performance of the system of interest is reviewed and its past and possible vulnerabilities are identified. Often past experiences with extreme hydrologic events offer the strongest evidence of the hazards that should be addressed. Each water-resource system is unique and has exposure to events of particular characteristics. For example, in a general way Vogel et al. (1997) classified reservoir systems into over-year or under-year storage categories based on their vulnerability to drought durations of months but less than years. Such systems would be interested in very different definitions of drought. It is also useful to consider the past experiences of other water-resource systems, especially those with similar size, similar climate, or similar operating constraints such as those posed by multiuse water systems and environmental requirements. Once the events of concern have been identified, the next part of this step is the characterization of the hazard through the estimation of probabilities associated with it. For waterresource systems, the events of interest are typically related to stream flow and so a hydrologic record that relates to the water system is genrally used to estimate these probabilities. Thus, a hazard is an event of interest (typically due to its potential to cause danger) with an associated probability of occurrence. Here, we focus on hydrologic hazards but the methodology is consistent with other hazards such as earthquakes or terrorist attacks. Note that the methodology also applies to events with positive consequences. These events opportunities are termed an event of interest (with potential for benefit) and its associated probability of occurrence. In the case of floods, the process is one of estimating the exceedance probability of certain high-flow events that were deemed to be dangerous. Similarly, the exceedance probability of droughts of specific duration and intensity (e.g., 20% reduction of inflows over a 6-month period) can be estimated. Traditionally, these probabilities were estimated using the historical record. For example, for floods, the annual maximum time series for the location of interest was used to estimate the parameters of an extreme value probability distribution. This distribution was then used to calculate the estimated exceedance probability for a flood of a given magnitude. The techniques for calculating parameters, choosing distributions, and using the distributions have been the focus of much attention and are fairly mature. A useful summary is found in Stedinger et al. (1993). Methods are also available for using data from other locations in a region when the location of interest has a limited historical record. For droughts, the process is similar, with the added dimension of the duration as well as the intensity of the event. Engineers have long been aware of the uncertainties that accompany the estimation of rare events such as flood events with low exceedance probabilities. These are addressed through risk management and communication. Nonetheless,

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a great deal of effort is expended in reducing the uncertainty of hazard estimates. This has yielded better understanding of the physical causes of hydrologic extremes and in some cases identified temporal structure in their occurrence. Our growing knowledge of the physical causes and temporal structure of the frequency domain of hydrologic extremes has important implications for the estimation of hazard probabilities. This leads to significant implications for our management of risks as well and perhaps the need for significant reconsideration of how climate related risk in water-resource systems is addressed. The most prominent example is the case of the El Nino/ Southern Oscillation (ENSO). ENSO is a coupling of seasurface temperature anomalies, winds, and atmospheric pressure in the equatorial Pacific that influences temperature and precipitation patterns around the world through teleconnections, with the largest effect in the tropics. The ENSO phenomenon has two phases, one warm (El Nino) and one cold (La Nina), each with partly but not entirely opposite directions of influence (warmer vs. colder; dryer vs. wetter). ENSO is described as quasi-periodic and these events occur on an irregular cycle of about 3–5 years. Where the teleconnections are strong, an ENSO event implies that the hazard associated with a particular hydrologic event may be elevated in some locations and reduced in others, relative to the longterm mean rate of occurrence. For example, the drought hazard in the Philippines is elevated during an El Nino, while it is reduced during a La Nina. The presence of such temporal structure has implications for risk assessment and management, since foreknowledge of year-to-year changes in risk was not previously considered.

1.11.2.2 Risk Assessment In this step the concept of risk is quantified in terms of the consequences of an event and its probability of occurrence, in accordance with Equation (1). Here the description of consequences and the specification of the consequence function are central. A challenge in calculating the consequences is that they are not necessarily measured in commensurate units. Many consequences are fairly straightforward to quantify in economic terms, and methods for doing so are mature. In the field of water resources, however, in many cases damages may be more difficult to describe in monetary terms. For example, flood events often involve the potential loss of life. Floods and droughts may also have substantial impact on the environment through the destruction of habitat or insufficient stream flows for aquatic ecosystems. This issue is more prominent in the risk management step, where the alternatives for managing hydrologic risks are evaluated in terms of their costs and the benefits provided and it is addressed further below. As before, the involvement of stakeholders is a critical aspect of risk assessment. Stakeholders are knowledgeable about the consequences of historical events and may be the best source of conjecture on the impact of events not yet seen. Their opinions will also be valuable in the ranking of risks that cannot be easily quantified, such as specific environmental

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damages or the relative value of structural damages in comparison with lost life.

1.11.2.3 Risk Management The final step in this process entails making decisions and systematically exploring the consequences of those decisions as well as the uncertainties that will accompany any risk assessment process. There are a variety of approaches that have been proposed for risk management. Here, we present risk management as a systematic process that can be described in three parts: (1) systematic evaluation of the risks previously quantified and the cost of measures that address those risks; (2) development of a plan to address those risks and accompanying uncertainty; and (3) consideration and planning for residual risks and performance of the system when failure occurs. The systematic evaluation of risks begins with a review of the risks quantified during hazard characterization and risk assessment. These risks can then be evaluated in terms of the expected losses that they represent. As described above, the expected losses may be expressed in economic terms, such as monetary units of damage to structures, or in units of human injuries and deaths, or in terms of damage to the environment. Since these expected losses are compared with the economic costs of the measures considered for mitigating the risks, expression in terms of economic units simplifies the development of a risk management plan. Methods are available for calculating the economic costs of injuries, deaths, and environmental damage, although none are without controversy. Environmental impacts are similarly difficult to quantify in economic terms, although methods have been proposed. At the least these potential impacts must be tracked and incorporated informally into the decision process. Next, the candidate measures for reducing risk are evaluated in terms of their cost and their effectiveness. Often, engineers limit themselves to consideration of structural approaches to managing hydrologic risks. For example, dikes and levees are typically prime candidates for reducing flood risk, whereas increasing the storage capacity of a water-supply system is the main method for addressing drought concerns. However, there exist a suite of nonstructural measures that should be considered as alternatives or complements for structural measures. For droughts, examples include economic mechanisms such as water markets, option contracts, and even insurance (Characklis et al., 2006; Brown and Carriquiry, 2007). For flood risk reduction, examples include designated flooding areas, improved flood warning systems, and land-use planning. An appealing aspect of some nonstructural measures is that they may be used as options, meaning that they are not executed or fully paid for until they are needed. For example, a levee can be heightened through the addition of sand bags which are not filled and put into place until they are needed (Lund, 2002). The option approach becomes more advantageous as the uncertainty related to the risk analysis increases, such as due to possible changing climate conditions. Also, the advantage of the option approach increases with improved forecasts that can result in better decisions related to

execution of the options. Unfortunately, our experience with these measures is limited as is the number from which to choose in comparison to traditional structural approaches. It is an area of needed research. When risks and the costs of risk reduction measures are well characterized, optimization and decision analysis can be an effective means for developing and evaluating the risk management plan. The objective function for a plan that considers permanent measures and option measures can be specified as follows:

min Z ¼

X þ

Z

permanent measures  cost ðoption measures  costs þ damagesÞf ðxÞdx ð4Þ

An example of the use of optimization using a form of this equation to develop a flood risk reduction plan is described in Lund (2002). Optimization techniques such as used in that analysis allow a quantitative approach to the development of a risk management plan. However, the result of that approach, say the optimal plan, is only as useful as the quality of the data that was used to develop the plan. In the case of estimating probabilities of rare events or equivalently the magnitude of events with very low exceedance probabilities, the quality of these estimations may not be very reliable. Furthermore, according to the engineer, the process of going from the optimal plan to a plan that is actually implemented by a public agency with diverse and conflicted stakeholders is quite difficult and not well understood. The latter subject may be the most important to solve in order to achieve efficient reduction of hydrologic risks to society. The subject is explored in Fiering and Matalas (1993). The uncertainty related to the estimation of magnitudes of rare events can be addressed in several ways during the development of the risk management plan. The first approach is simply the application of decision analysis and should be a standard part of the risk management plan development. Using a decision analysis approach, it might be determined that the decision is not sensitive to specification of these very rare events. For example, it may become clear that a particular flood risk reduction candidate measure dominates other measures regardless of the magnitude of floods with exceedance probabilities of 103 or 104. Or the decision to provide protection for a structure may be decided based upon what a society is willing to pay rather than the estimated return period of an event that is protected against (Bondi (1985) cited in Plate (2004)). In other cases, the choice of a given plan and its subsequent performance is sensitive to these estimations. In such cases where uncertainties are significant and influential, a risk management plan should strive to achieve aspects of robustness and flexibility (de Neufville, 2004). While it can be defined in a variety of ways, here we define as robust the quality of a system that can continue to operate effectively under unexpected conditions. Redundancy is a key component of achieving robustness. For example, two pumps should allow water to be moved even if one unexpectedly fails. A groundwater well field allows a water-supply system to

Risk Assessment, Risk Management, and Communication: Methods for Climate Variability and Change

continue to operate if the reservoir runs dry. The quality of flexibility is a quality of a water system that provides the engineer or manager with options to react to changing conditions as they evolve. For example, a regulation plan for a reservoir could allow different release rules based on the prevailing climate conditions in place of the standard static regulation plan that is unchanging. Brown et al. described the development of a dynamic regulation plan for the management of outflows from Lake Superior that demonstrates how instilling flexibility in normally static plans is an adaptation strategy for climate change. Flexibility also allows the water manager to make use of forecast information. With only static operating procedures, foresight regarding future water supplies or demand is of little value since there are no changes that can be made in response to the new information. Once a risk reduction design has been formulated, the final step consists of addressing the full spectrum of possible risks that a system faces. There are two aspects to this. The first aspect is the consideration of residual risk. Residual risk is the risk that remains once a particular decision has been made. The concept applies to operational or planning timescales. For example, a planned flood risk reduction design may include the construction of a levee that is designed to withstand a flood magnitude with an estimated 500 year return interval. The residual risk consists of the risk that is not addressed by the design, in this case floods that are greater than the 500year flood. Such floods are certainly possible and must be planned for even if they are not part of the structural design considerations. On an operational timescale, it is the risk that remains once a specific operational design has been made under uncertainty. The residual risk for decision i may be specified as

RRi ¼

Z

N

CðxjDi Þf ðxÞdx

ð5Þ

0

The residual risk may be addressed in a variety of ways. It may be considered insignificant and no further action is taken. A decision analysis approach could be used. Using this method, the costs of addressing the residual risk could be estimated and compared with the risk. In most cases the residual risk estimate will be highly uncertain since by nature it involves the tails of distributions. The decision will be largely a subjective one and may or may not be influenced by the precautionary principle. Since in most cases it will not be economically acceptable to address all residual risks, there will be risks that are known but are not addressed. There will also be risks to the system that have been underestimated. The uncertainty associated with climate change, land-use change, and technological change compromises our ability to estimate future probabilities with confidence. As a result, we should not be surprised by underestimation and overestimation of risk. We should be surprised by risks that are unknown and cannot be reasonably anticipated. These are surprises or even black swans as popularized by Taleb (2007). From the water engineer standpoint, these three considerations make clear that any system design that incorporated risk reduction planning may still fail. This brings us to the second aspect of addressing the full spectrum of risk. Although risk reduction plans may

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not cover all risks, we can plan for how our system operates during an inevitable failure. This may seem to fall outside the traditional realm of engineering. But since the engineer is ultimately trying to reduce impacts of hydrologic events on society (or rather reduce negative impacts and increase positive impacts), this represents more options for doing so successfully. For a water-supply system, a plan can be developed for rationing water, purchasing tanker water deliveries, and other measures if a drought were to shut down the reservoir supply. Similarly, emergency response, evacuation planning, designated flood areas, and secondary protections beyond primary levees should be incorporated into a flood risk management plan. The key is for managing surprises such that although they lead to failure of the system, they do not lead to catastrophe.

1.11.3 Risk Management versus Consequence Management: The Upside of Risk Risk management focuses on uncertainties related to events that have the potential to cause damages or harm. These are typically events that differ from the average conditions. Therefore, droughts are brought on abnormally by low water supply availability and floods by water that exceeds normal conditions. In some cases or for some stakeholders, these relatively rare events represent opportunities instead of risks. For example, excess water represents the opportunity to generate additional energy for a hydroelectricity generator or to provide high flow for ecosystems that require it. Droughts may provide opportunities for offline inspection and maintenance work. The concept of planning for the opportunities as well as for the risks associated with uncertainty has been described as uncertainty management (de Neufville, 2004). Here, the term consequence management is preferred as this is a neutral term of risk (positive or negative) and is distinct from uncertainty which derives from a variety of sources and does not necessarily have consequences. With consequence management, we include the positive consequences of possible hydrologic conditions in addition to the negative consequences (risks). Management plans are developed and evaluated for all consequences in the same way as done for risks alone. Interestingly, the management of risk (negative consequences) and the ability to exploit opportunities (positive consequences) are connected in many cases. The reason is that if risks are not being effectively managed, it is more difficult to capitalize on the opportunities associated with favorable side of the distribution. As a result, there can be substantial lost opportunities that result from ineffective risk management. Examples from the use of seasonal climate forecasts demonstrate this issue. With the forecast for above-average reservoir inflows, a multi-use reservoir can consider making additional releases for hydroelectricity production while still expecting to meet its storage requirement for water-supply purposes. However, the typical water manager would prefer not to make the additional releases because of the concern that belowaverage inflows may occur even if the probability is very low. As a result, the excess water and the opportunity it represents are lost as spills from a full reservoir. This decision is perfectly rational if the consequences associated with a shortage of

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water in the reservoir are disastrous even if the odds are low. Without an effective plan for managing low water conditions the risk associated with additional reservoir releases outweigh the benefit of extra hydroelectricity production. If there is an effective way to manage the downside risk, then the upside consequences could be better captured. So a system that has a backup water supply or the means to buy-out major water users (such as agriculture) is in a good position to take advantage of such an opportunity. These issues are explored in a case study based on the multipurpose reservoir system that supplies water to the city of Manila, the Philippines (Brown et al., 2008; Brown and Carriquiry, 2007).

1.11.4 Climate Risk The subject of climate change and nonstationarity has stirred interest in risk-based approaches within water resources planning and management. Specifically, there is an emerging question as to whether our traditional approaches to assessing and managing risk are appropriate with the recognition of nonstationarity. While there is naturally strong interest in improving our ability to estimate the return periods or exceedance probabilities of hazardous events, there is also a cause to reconsider more generally how we assess and manage risk. Initially, the sense of the water community was that there was too much uncertainty to take much action in response to climate change. The resulting focus emphasized the use of output from general circulation models (GCSMs) to reduce that uncertainty. This style of analysis typically utilized a scenario approach whereby a few simulations of possible future conditions were generated from a particular GCM run representing a possible future or scenario. The approach generally consisted of the following steps (see, e.g., Brekke et al., 2009): (1) choose a GCM or a few GCMs from which to derive the climate change model projection; (2) downscale the coarse-scale GCM output to the scale of a higher-resolution hydrologic model; (3) use the hydrologic model to produce estimates of stream flow for the climate change scenario; and (4) use the stream flow estimates to estimate a water resources impact of interest, such as reliability of a reservoir system, change in production of hydroelectricity, and implications for low stream flows. While seemingly straightforward, there are a number of drawbacks to this approach when used for planning and decision making. There are a variety of difficulties associated with the use of the GCM information itself. Typically mean conditions from the GCM are utilized. As a result, climate change is depicted as a shift in mean conditions. However, the water resources sector is usually more concerned in variability and extremes. For the water-supply sector, changes in the serial correlation of rainfall or stream flow are of concern, since several months or years of continuous below-normal conditions are the greatest risks. For risk reduction, extreme hydrologic events are more important than the mean conditions. The scenario approach does not attempt to address changes in variability or in the distribution of extremes. A more fundamental critique of this approach is that it is not risk based. The approach produces possible consequences that result from a particular scenario of climate change. There

are no probabilities associated with the particular scenarios or the consequences. As a result, there was no way to quantify the risk that climate change posed. The focus of research efforts turned to attempts to improve the ability to simulate future climate conditions. Emphasis was placed on increasing the resolution of the climate models and incorporating more and better understanding of the processes that comprise the Earth’s climate system. While this emphasis continues, the difficulty in reducing the uncertainty significantly due to the complexity of the Earth’s climate system and predicting the evolution of civilization over the next 100 years has led to dissatisfaction with this emphasis for use in water resources planning. On the topic of climate change, until recently there has been more focus on risk-based approaches within the adaptation segment of the climate change community generally than within water. This is reflected in the Intergovernmental Panel on Climate Change (IPCC) reports and broadly in the literature. While there are a wide variety of approaches, one might place them in one of two general categories. The first is the top-down approach. In this approach, the analysis begins with some attempt to predict future conditions and those future conditions are used to estimate impacts on society, ecosystems, hydrologic systems, etc. This approach does not typically broach the subject of probabilities. In some sense they may not be truly risk based for that reason. The notion of top down refers to the direction of analysis beginning from climate change impacts through various physical processes and finally considering the potentially impacted systems. The second category is characterized by the opposite direction of analysis and, for that reason, is referred to as a bottom-up or vulnerability-based approach. The analysis begins with the vulnerability, that is, consideration of how a system, such as a community, is susceptible to harmful effects from changes in climate. Thresholds can be set where systems are particularly vulnerable or where significant impacts begin to set in that are meant to be avoided. Since we define risk as a product of consequences and probability of those consequences, the estimation of probabilities of consequences is ostensibly a primary concern of risk management. We are severely limited in our ability to estimate the probability of future events. For many extreme events, this is true whether one considers climate change or not. We have very few observations of very rare events that the magnitude of particular percentile events is a function of the data which consists primarily of much more common events (of lower magnitude) and the choice of the extreme value distribution used to model them. A review of the methodologies explored in the estimation of probabilities of extreme events such as floods is beyond the scope of this chapter. The reader is referred to Stedinger et al. (1993). Here, it is noted that the prospects of getting the answer correct seem to be slim. From a risk perspective, a risk that must be managed is the very real risk that we are unable to accurately estimate the magnitude of a design flood or drought.

1.11.4.1 Risk and Nonstationarity: Uncertain Information and Unreliable Probability Estimates The prospect of climate change has raised to new prominence the specter that our estimates of probabilities associated with

Risk Assessment, Risk Management, and Communication: Methods for Climate Variability and Change

hydrologic events based on historical records may not be reliable in the future. Simultaneously, the projections of future climate conditions from GCMs are not reliable enough to be used to replace the historical record (Brown et al., 2009). As a result, water engineers face a future which has more uncertainty than they have previously been aware. Risk management is an applicable framework for addressing these issues. The methodology described here still relies on estimates of probabilities that are most likely to come from the historical record. Due to uncertainties related to climate change, among others, the final step of risk management, especially consideration of residual risk, surprise, and operation in failure mode, is essential. In addition, methods for decision making under so-called deep uncertainty may be explored. Robust decision making is representative of these methodologies (Lempert and Collins, 2003). The approach begins from a traditional statistical decision analysis approach but then evaluates decisions free of probability estimation. Instead, a candidate decision is evaluated for the conditions of the state variables under which that decision is compromised or performs poorly. Cluster analysis is used to identify the major areas of concern. Trade-off analysis is then used to evaluate alternatives to the original candidate that perform in these areas of concern. The approach appears computationally intensive in practice but the authors report good results in a few case studies (Lempert et al., 2003). Robust decision making is useful when little can be said of future conditions. However, there is some insight available regarding climate change. It may be possible to utilize the decision analysis if the process is tailored for the specific issues related to the use of uncertain information from GCM and historical data. Such an approach is presented here. The approach is called climate consequence management to highlight the possible opportunities that climate variability and change may present.

1.11.4.2 Climate Consequence Management with Decision Scaling: An Approach Designed for Uncertain Information Climate consequence management is a version of risk management that has been designed for the specific case of uncertainties associated with climate variability and change. The fundamental breakthrough is that a decision-analytic approach is employed in a manner that allows the use of climate information in a bottom-up or vulnerability-based way. It emphasizes identification of the specific climate information required that a decision hinges on. This allows the tailoring of the climate information to fulfill decision needs. The premise is that climate information can be useful in certain conditions by setting the bar fairly low in terms of what is needed. In some cases, for example, where climate information is deemed fairly reliable and projections are consistent in direction, this allows for probabilistic estimates of risk and risk-weighted decision making. There will be other cases where projections are contradictory and the process enables the identification of climate sensitivities and provides a framework for addressing them. In other cases, it will be found that some decisions are not sensitive to the projections of

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climate change due to limited options or limited effects. The term ‘consequence’ is used to highlight the concept that variations and changes in climate bring opportunities as well as risks. The process consists of three steps that are analogous to the three-step process described above. In the first step, vulnerability and uncertainty identification, the hazards, and opportunities associated with climate are identified. Next, consequence assessment is the process of estimating probable consequences from the hazards and opportunities identified in step 1 and tailored climate information produced through the process of decision scaling to assign probabilities to the hazards and opportunities. In the final step, consequence and uncertainty management, a strategy for addressing probable consequences and key uncertainties, is being developed using a decision analytic framework. Decision scaling is again employed to tailor climate information to aid in the analysis of different decision options. Finally, residual risk and surprise are addressed through the incorporation of robust and resilient design, which may in some cases include adaptive management.

1.11.4.2.1 Step 1: Vulnerability and uncertainty identification The first step of climate consequence management is to assess the impacts of changes in climate across all timescales on water resources. This necessitates knowledge of both historical climate information and the resulting local consequences. The historical record is a useful starting point for identifying how climate has impacted the system in the past and the particular climate episodes that are challenging. A general overview of climate change information for the region being studied is accessed to prompt consideration of potential climate impacts that have not been observed in the historical record. At present, the IPCC reports on regional impacts are logical starting point for a summary of literature. The brainstorming process of identifying future potential impacts is not limited to these projections since they are not certain to describe the full range of future climate possibilities. It can also be very useful to identify thresholds in the system performance that when exceeded signify the need for adaptive actions. These thresholds are used in the consequence assessment and consequence management steps that follow to determine whether action is necessary in response to projected risks. Developing the appropriate knowledge requires a dialog with the stakeholders affected by or engaged in the water sector. Engaging stakeholders can both ensure that relevant impacts are considered and keep stakeholders aware of the process. A dialog with climate scientists and meteorological agencies can help supplement and interpret relevant climate information. By gaining a more robust understanding of these hazards and impacts, one can begin to determine the hydroclimatic risk and opportunity for a given system. While the focus naturally turns to the negative climate events in the historical record, it is important to consider also the effect that uncertainty in general has on the system. For example, preparations for a rare negative event may cause considerable lost opportunities in the many years the event does not occur. This can represent a significant opportunity cost due to

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uncertainty. Better management of the vulnerability to the rare event may allow the exploitation of more opportunities. While this process focuses on climate risks, it is important to recognize that climate is one of many factors affecting the system. When projecting future risk scenarios for a given system, possible changes in population growth, water demand, land use, and even values should also be considered and integrated into any comprehensive risk assessment. The following questions provide a general guideline for what to consider when performing vulnerability and uncertainty identification. What key climate-related challenges does the system currently face? These challenges might include moderate or severe droughts, flood events, variable flows, or others that are particularly disruptive to the system. This assessment is based on historical climate and current system characteristics, such as land use, population, and economic factors. It is important to identify the hazards historically associated with climate variability for the system while understanding also that the same type of climate event might have a more or less severe impact based on evolving nonclimate characteristics of the system. What damages occur as functions of these events? Once one has identified the climate-related hazards, one should assess the local impacts on the system. This includes an analysis of the spatial distribution of impacts and a determination of whether there are distributional effects from these events. Distributional effects cause some populations, such as those at lower economic status, to be more vulnerable to hazards (Rees, 2002). One should also determine impacts on both the human and environmental systems and seek to assess their vulnerabilities. The method of valuing consequences may differ. For example, economic valuation of consequences (e.g., foregone profits, direct costs associated with switching to another water source) is appropriate in some cases. However, in the case of severe consequences (e.g., famine), economic valuation alone may not be sufficient, as the social consequences may far outweigh direct economic costs. It may be important to determine local thresholds that determine the extent of climate-related consequences. While some water users can easily adapt to small reductions in water supply with little or no adverse effects, others may face significant damages from even the smallest supply variations. This may also be true for floods. The vulnerability across different users might lead to an aggregate threshold level and expected reliability for the system. Are there potential opportunities due to climate variability and change? Although the emphasis is generally on the possible negative impacts from climate variability and long-term change, changes in the system might also bring benefits. For example, a shift in phase in multi-decadal variability within a system could lead to improved average climate conditions for some sectors. Consider the apparent upswing in West African rainfall recently. It is important to remember interactions between climate variability and the possible impact of long-term climate change. The latter might also offer some opportunities (e.g., increased average precipitation in arid regions). Assessments should take into account the varying opportunities and risks across sectors and across or within regions, along with their uncertainties.

Are there opportunity losses due to decisions made to avoid current climate risks? Water-resources managers are typically quite risk averse, meaning that they would prefer an option with less uncertainty but possibly a lower net benefit over an option with greater uncertainty but a higher possible net benefit. Thus, decisions that minimize climate risks may also decrease the potential benefit and result in lost opportunities. Identifying these lost opportunities reveals increased possible benefits from improved climate consequence management. Have the occurrences of hazard events over the historical record followed identifiable patterns? The initial step is to determine recurrence periods for relevant climate events over the historical record. For example, the analysis might reveal how frequently the system has experienced severe droughts. In some cases, there is spatial or temporal structure (i.e., a pattern) in the historical hazard occurrence. This might include variability across various timescales (such as inter-annual variability due to ENSO) or longer-term trends. The main purpose at this point is to understand variability in the climate system and how it has affected hazard probabilities in the past. One is not yet making forecasts or projections about future scenarios. This analysis reveals the probabilities that have determined system risk up to the current period. The understanding of historical climate variability at different timescales also suggests the key components to consider in developing projections in future steps. This may include identifying appropriate predictors that can help one make simple forecasts of possible shifts in the probability distribution of supply in the system (e.g., shifts due to ENSO phases). Can thresholds be identified that represent changes in system performance that require action? Hydro-climatic conditions affect a water system’s ability to meet performance objectives. Climate variability and change have a significant impact on whether the system fails or is able to meet stakeholder needs. Given the uncertainty related to climate change, it is useful for decision purposes to identify thresholds that signal where a system performance is no longer acceptable and adaptive action is necessary. For example, a reservoir may be designed for a long-term reliability of water delivery of no less than 95%. A threshold of 95% reliability would then be appropriate, for if the reliability fell below this level action would need to be taken. Analysis and answers to the previous questions in this section provide data on historical climate variability and probabilities associated with various climate outcomes. If climate conditions and the historical variability were expected to continue into the future without any changes, one could model the expected reliability based on past experiences. However, this assumes that one is aware of all forms of variability in the past and has the ability to model the future with a high degree of accuracy. If the historical record is too short to capture the full range of climate variability (and this is not uncommon), the results of the analysis can be significantly biased due to sampling variability. In addition, this does not take into account the possible nonstationarity of the system. In order to address these concerns and appropriately assess the sensitivity of the system, it is best to model system performance based on both historical data and scenarios of

Risk Assessment, Risk Management, and Communication: Methods for Climate Variability and Change

possible future climate conditions. A procedure for doing so, decision scaling, is described below.

1.11.4.2.2 Step 2: Consequence assessment This step corresponds to the risk assessment described earlier. In the general sense, this is the product of the probability of some event and the consequences of the event. The process described here is designed for the special case of using climate information for the estimation of probable consequences. This process is termed decision scaling and described below. Decision scaling: tailoring climate information for risk assessment. Decision scaling is the process of tailoring climate information to aid in decision making related to climate risks and opportunities. It uses a decision analytic framework to identify the needed climate information, which is then produced through tailoring climate information from all possible, reliable sources. In general, this means that first the decision must be identified and next the points at which the decision changes as a function of climate information are specified. By doing so, the specific climate information can be tailored from various sources, including GCMs and paleoclimatological data. For example, if it is known that a reservoir system is vulnerable to multi-year droughts and inter-annual variability, but not vulnerable to within year variability, then the effort to produce climate information focuses on trends and temporal structure in inter-annual variability. The premise of this approach is that there are significant and irreducible uncertainties associated with projections of future climate and the resulting hydrologic conditions. Therefore, the emphasis should shift away from attempting to predict the future or provide a few examples of possible future climates. The usual process of analysis starts with GCMs and all their uncertainties, a few scenarios are generated and then those are input to a system model to review the impacts. In decision scaling, the process begins with the system or decision model which is perturbed with parametrically varied climate to generate a climate response function, that is, a representation of the systems’ sensitivity to climate changes. Then the GCM or other climate data are used to estimate the subjective probability of the climates of interest – these climate changes that cause significant impacts, exceed thresholds identified in step 1, or are related to a point where a decision changes. The most effective way to do this is through the use of multiple run, multi model ensembles to utilize the fullest representation of the model projection uncertainty. Decision scaling is used during the consequence assessment to describe probable consequences or the probability of exceeded adaptation action thresholds. It is used again during Step 3: Consequence and uncertainty management to aid in the decision making process. Also, the focus of recent interest in risk assessment and management is largely related to climate change implications. However, the process described here is appropriate for climate variability as well. Some particular considerations related to climate variability are described below. Consider uncertainty in climate forecasts. Based on location and climate characteristics, there may be significant variation in the ability to make climate predictions. For the same system, forecast skill might vary significantly across time scales.

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It is critical to be aware of the predictive capacity for the given system and the uncertainty associated with any predictions. The probabilistic nature of climate forecasts reinforces the idea that they are neither guaranteed nor absolute. This uncertainty plays a significant role when integrating the climate information into decision making and one should explicitly assess the uncertainty of any forecasts you consult. The approach to assessing the forecast uncertainty depends on the techniques used to create the forecast and the projected time scale. For example, if a seasonal forecast has been developed using a statistical model, a cross-validation technique can be used to understand and quantify the uncertainty in the model. With complex dynamical and GCM-based models and projections over longer timescales, it is best to consult climate professionals to determine the uncertainty and errors present in the model. Some of the key discussion points regarding longer-term climate projections that include the effects of increasing greenhouse gases and other anthropogenic influences include 1. the climate model’s ability to reproduce climatology in the region; 2. whether the model captures the observed regional trend in twentieth century climate; 3. the extent to which there is a well-established physical basis for the model’s forecasts; 4. the degree of agreement between different models; and 5. the extent to which natural multi-decadal variability impacts the region. Asking these questions and validating forecast models can show where the model made errors and help understand that possible model weaknesses. This demonstrates the remaining uncertainty that must be addressed through management options, as discussed in the next step. In the case of climate change, uncertainty may be considered too significant to assign probabilities to specific climate events. Even when probabilities are assigned, they must be considered fairly broad estimations and due consideration given to surprise, resilient design, and performance in failure mode.

1.11.4.2.3 Step 3: Consequence and uncertainty management The final step consists of developing a plan to manage probable consequences and uncertainty that have been identified and described in the previous steps of this process. The process proceeds in the three parts described in Section 1.11.2. Decision scaling is used to provide climate information that is needed in the decision analytic framework used to evaluate alternatives. Here, decision scaling provides the probabilities of the scenarios that cause a particular option to be preferred over another. The estimated hydro-climatic risk determined in the previous two steps serves as the foundation for developing a portfolio of options to mitigate the risk and take advantage of possible opportunities. It is critical to realize that climate information provides information about the probability of particular climate events (such as droughts), but anything can still happen, even if it is very unlikely. For this reason, we again highlight critical aspects of consequence and uncertainty

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management that are vital to managing risks and opportunities associated with climate variability and change. These are residual risk, surprise, and performance of the system when failure occurs. In the context of climate uncertainty, systematically evaluating risks and opportunities due to climate and comparisons with costs of various actions will benefit from a decision analytic framework. Decision scaling facilitates the production of relevant climate information for use in such a framework. For example, decision analysis might determine that decision A is optimal for a future in which mean annual stream flow decreases by 10% or more, while decision B is optimal in all other cases. Decision scaling is used to estimate the probability of decision A being the better choice using the full range of climate data available. While there is uncertainty related to that probability, if it is a large number relative to the probability associated with the optimality of B, say more than 60%, then we have some confidence that A is the right choice even with the uncertainty. Through such analyses, a plan is developed for addressing risk, opportunities, and uncertainty. The degree of uncertainty associated with future climate variability and change may favor delay in decisions where appropriate. It may also favor the methods for managing the impacts of climate events that do not necessitate new investments in infrastructure. The reasoning is as follows: if an event is not very likely to occur, it is typically not worth making major investments to manage the impact. However, given the possibility that we have underestimated a risk, we should consider ways to avoid the negative impacts of that event, if possible. Finding solutions that can be called upon only when needed is an efficient way to manage the impacts of unlikely events. Residual risk and surprise may also be addressed through redundancy. If a water supply system consists of a single source, any impact on that source leaves the system vulnerable. While it may not be economically efficient to build new infrastructure to tap new sources, other opportunities may exist. The suite of risk management options might include economic instruments (such as insurance or water banks), infrastructure modifications, or integrating seasonal forecasts into decision making, among many others. Together, these approaches are termed portfolio of options because they consist not of a single solution, but rather a basket of possibilities – each of which may be the best choice in a particular circumstance. The final consideration is the ability of the system to perform when the primary risk management approaches fail. For a water-supply system this situation may occur due to the loss of the supply due to drought. For flood protection this is the overtopping of the levees. While the primary system has failed this does not mean the plan has failed and should not mean catastrophe. Through residual risk management and redundancies as described above, the system may be able to continue performing. Water-supply systems may use boil water orders or may contract with neighbor utilities for trucked water. In floods, the planning of evacuations and local emergency response can ensure that an extreme hydrologic event does not have extreme human impact.

Below are some additional considerations when developing the portfolio and determining the most appropriate solutions. Consider planning and operational approaches. The risk management solutions available depend partly on the time frame for action. Near-term operational options will most likely assume fixed infrastructure and some level of sunk costs (those that have already been allocated and cannot be recovered). Possible planning solutions, on the other hand, can include decisions regarding infrastructure and system design. Climate information should be integrated into decision making at the appropriate time scale to inform options most effectively. Projections of long-term climate change may have little value at the operational level for current practices. However, such projections might inform planning decisions as well as the framework under which operational decisions are made in the future (i.e., whether expected climate changes necessitate more flexible operational policies). Assess possible trade-offs. Limited human, financial, and natural resources lead to trade-offs in almost all decisions in water resources management. Water managers must seek to understand and assess possible benefits or consequences of their decisions within the context of these resource constraints. Uncertainty makes such assessment even more difficult, but can also increase the importance of decision outcomes. There is often a trade-off between increasing expected reliability for a system and increasing possible benefits from water allocation. Improved climate information and projections of likely futures may help shift the reliability scenarios. While this does not eliminate the necessary tradeoff, it can improve the possibility of achieving positive outcomes. Integrating thresholds of acceptable costs into decision making can help water managers balance trade-offs. Consider the impact of uncertainty. It is necessary to understand the uncertain nature of probabilistic forecasts in order to assess the suite of options appropriately. Rather than planning for a specific outcome, the most appropriate approach often requires planning for a set of scenarios. While the likelihood of a specific outcome might be higher than the likelihood of another, both are possible and should be considered in decision making. This uncertainty may lead to more flexible approaches and policies, with less emphasis on rigid options that leave little room for alternative outcomes. A flexible, adaptive plan might also increase the capacity to take advantage of possible opportunities from better than expected outcomes. Of particular importance is to consider the effects of low-probability but high-impact events on the system when actions are taken based on a forecast. For example, if the forecast leads one to expect more water, are there ways to mitigate the effects of an unlikely severe drought? This is important to consider because sometimes the anticipatory actions based on a forecast may leave a system more exposed to the down-side risk, or the risk associated with the less likely, but still possible, climate extreme.

1.11.5 Conclusion In this chapter an approach to risk assessment, management, and communication is presented that attempts to reconcile

Risk Assessment, Risk Management, and Communication: Methods for Climate Variability and Change

traditional approaches with our growing knowledge of uncertainty and nonstationarity that mark the hydrologic record. In describing these steps, it becomes clear to the authors at least that the research emphasis within the water community has focused primarily on the means to reduce the uncertainty related to hydrologic events and better prescribe the distributions used to estimate their probabilities. Relatively little effort has been devoted to develop innovative means of reducing hydrologic risk to society or of communicating risk in order to promote risk-reducing behavior. The effects on the hydrologic record of climate change and land-use change calls this research orientation into question. Irreducible uncertainties that hamper our ability to estimate hydrologic design variables imply that greater effort is needed for the development of designs and strategies that perform well under a range of possible future conditions. By developing a wider range of options, water engineers and planners will have greater opportunity to design systems that are resilient as conditions evolve into the future. To develop these options, new ideas related to the use of information technology, emergency planning, and economic incentives should be explored. The challenge of managing risks which cannot be effectively constrained with our traditional approaches means we must be willing to expand beyond our comfortable set of tools. If we are not willing to manage the full range of hydrologic risks, who do we expect will?

References Bondi H (1985) Risk in perspective. In: Cooper, MG (ed.) Risk: Man-Made Hazards to Man. New York: Oxford University Press. Brekke LD, Kiang JE, Olsen JR, et al. (2009) Climate Change and Water Resources Management – a Federal Perspective, U.S. Geological Survey Circular 1331, 65p (ISBN 978-1-4113-2325-4). http://pubs.usgs.gov/circ/1331 (accessed April 2010). Brown C and Carriquiry M (2007) Managing hydroclimatological risk to water supply with option contracts and reservoir index insurance. Water Resource Research 43: W11423 (doi:10.1029/2007WR006093).

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Brown C, Conrad E, Sankarasubramanian A, and Someshwar S (2009) The use of seasonal climate forecasts within a shared reservoir system: The case of Angat reservoir, Philippines. In: Ludwig F, Kabat P, van Schaik H, and van der Valk M (eds.) Climate Change Adaptation in the Water Sector, 274pp. London: Earthscan. Brown C, Greene A, Block P, and Giannini A (2008) Review of downscaling methodologies for Africa applications. IRI Technical Report 08-05, 31 pp. Characklis GW, Kirsch BR, Ramsey J, Dillard KEM, and Kelley CT (2006) Developing portfolios of water supply transfers. Water Resources Research 42: W05403 (doi:10.1029/2005WR004424). de Neufville R (2004) Uncertainty Management for Engineering Systems Planning and Design, Engineering Systems Monograph. p. 18. Cambridge, MA: MIT Press. Dessai S, Hulme M, Lempert R, and Pielke R, Jr. (2009) Do we need better predictions to adapt to a changing climate? EOS Transactions AGU 90(13) (doi:10.1029/ 2009EO130003). Fiering MB and Matalas NC (1990) Decision-making under uncertainty. In: Waggoner PE (ed.) Climate Change and U.S. Water Resources, pp. 75–83. New York: John Wiley & Sons. International Research Institute for Climate and Society (IRI) (2006) A gap analysis for the implementation of the Global Climate Observing System Programme in Africa. IRI Technical Report Number IRI-TR/06/1. New York: IRI. Lempert R and Collins M (2007) Managing the risk of uncertain threshold responses: Comparison of robust, optimum, and precautionary approaches. Risk Analysis 27(4): 1009–1026. Lund JR (2002) Floodplain planning with risk-based optimization. Journal of Water Resources Planning and Management 3: 202–207. Plate EJ (2004) Risk management for hydraulic systems under hydrological loads. In: Bogardi JJ and Kundzewicz ZW (eds.) Risk, Reliability, Uncertainty and Robustness of Water Resources Systems, 220pp. New York, NY: UNESCO and Cambridge University Press. Plate EJ (2002) Risk management for hydraulic systems under hydrological loads. In: Bogardi JJ and Kundzewicz ZW (eds.) Risk, Reliability, Uncertainty and Robustness of Water Resources Systems, chap. 23, pp. 209–220. Cambridge, UK: Cambridge University Press. Rees J (2002) Risk and Integrated Water Resources Management. Global Water Partnership TEC Background Paper No. 6. Stedinger JR, Vogel RM, and Foufoula-Georgiou E (1993) Frequency analysis of extreme events In: Maidment DR (ed.) Handbook of Hydrology, chap. 18. New York: McGraw-Hill. Taleb NN (2007) The black swan: The impact of the highly improbable. New York: Random House, 400pp. UNDRO Office of the United Nations Disaster Relief Coordinator (1991) Mitigating Natural Disasters: Phenomena, Effects and Options. A Manual for Policy Makers and Planners. New York: United Nations. van Aalst M, Hellmuth M, and Ponzi D (2007) Come rain or shine: Integrating climate risk management into African development bank operations, Working Paper No. 89. Tunis: African Development Bank.

Preface – The Science of Hydrology S Uhlenbrook, Department of Water Engineering, DA Delft, The Netherlands & 2011 Elsevier B.V. All rights reserved.

The world is changing and it seems that the speed of changes is accelerating. In the overall introduction of the Treatise on Water Sciences, the editor-in-chief Peter Wilderer (The Importance of Water Science in a World of Rapid Change: A Preface to the Treatise on Water Science) discusses the prevailing changes, its drivers, and possible impacts on different water disciplines. A major challenge is that all changes and their various impacts are interacting with each other, although how and to what extent is often poorly understood. For scientists and practitioners, this makes the problem identification and the development of sustainable solutions for water problems a very difficult task. Therefore, it is very timely to summarize the contemporary state of the knowledge in the different fields of water sciences and technology, and to provide a platform for innovative research and development. I am pleased to conclude that this volume on hydrology is an important piece of the complex puzzle. What is hydrology? The International Association of Hydrological Sciences (IAHS) in collaboration with UNESCO defined hydrology as the ‘‘science that deals with the water of the earth, their occurrence, circulation and distribution, their chemical and physical properties, and their reaction with their environment, including their relation to living beings.’’ In addition, it states that hydrology is the ‘‘science that deals with the processes governing the depletion and replenishment of the water resources of the land areas of the earth, and various phases of the hydrological cycle.’’ This is indeed a very wide definition. Many aspects of the chemical properties and interactions with the environment are part of Volume 3 of this treatise. Topics that are directly related to the management of the water resources are part of (Preface – Management of Water Resources) of this treatise. However, this volume (The Science of Hydrology) deals with all major components of the water cycle and key water-quality aspects. It also discusses the linkages to closely related disciplines. The aims of the science of hydrology were well summarized by the Dutch Foresight Committee on Hydrological Science (KNAW, 2005) as follows:

1. to understand the mechanisms and underlying processes of the hydrological cycle and its interactions with the lithosphere, atmosphere, and biosphere; 2. to enhance our knowledge of interactions between the hydrosphere and atmosphere, the hydrosphere and lithosphere, and the hydrosphere and biosphere, thereby increasing our understanding of the role that water plays in the Earth system; 3. to quantify human impact on the past, present, and future conditions of hydrological systems; and 4. to develop strategies for sustainable use and protection of water resources, hydrological systems, and the associated environmental conditions.

The science of hydrology is special, as it holds a place, on the one hand, in the field of Earth System Sciences, where it is directly linked to earth science disciplines, such as atmospheric sciences, geomorphology, geology, soil sciences, geobiology, and ecology. On the other hand, hydrology is an applied science and, as such, a part of engineering. This makes the discipline highly relevant to the management and development of the water resources and the prediction and mitigation of water-related natural hazards (floods, droughts, landslides, etc.) to finally support life, civilization, and sustainable development. These complementary aspects of hydrology (Earth System Sciences and the basis for water management/engineering) make it an exciting and very relevant discipline. It is quite a dynamic discipline given the significant developments of the past decades; many of them are reviewed in this volume. The volume starts with a comprehensive overview of global hydrology and the spatio-temporal variability of hydrological fluxes and water resources on a large scale. It continues with several chapters on the main variables of the water balance, such as precipitation, evaporation and interception, and stream discharge; then it goes on to discuss the storage components of groundwater, soil water, lakes, and reservoirs. Unfortunately, a chapter on snow and ice, the globally largest and regionally/locally often very important water storage component, was withdrawn at a late stage and could not be replaced in time. The volume continues with several chapters discussing the state of the art and the possible future developments of observation methods for ground-based techniques (i.e., fieldbased methods, tracer techniques, and hydrogeophysics) and remote-sensing techniques. Key data analysis and modeling techniques as well as theoretical considerations are reviewed in four, mainly theoretical, chapters on scaling and regionalization, statistical methods, hydrological modeling, and uncertainty estimation techniques. The linkages between hydrology and aquatic ecology and biogeochemistry are discussed in two comprehensive chapters. Two chapters are related to the processes and issues of erosion and sedimentation as well as surface water–groundwater interactions. The inclusion of all these topics results in a sizable volume with 20 chapters, exceeding 500 pages. However, several hydrology-related topics are not or could be only partly covered (e.g., urban hydrology, snow and ice, coastal hydrological systems, landscape evolution, and hydrogeomorphology). Perhaps this can be seen as an invitation to redo the exercise in a few years from now, and to review the latest developments in this dynamic field and strive for more completeness.

References KNAW (2005) Turning the Water Wheel Inside Out. Foresight Study on Hydrological Science in the Netherlands. Amsterdam: Royal Academy of Arts and Sciences.

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2.01 Global Hydrology T Oki, The University of Tokyo, Tokyo, Japan & 2011 Elsevier B.V. All rights reserved.

2.01.1 Introduction 2.01.1.1 The Earth System and Water 2.01.1.2 Water Reserves, Fluxes, and Residence Time 2.01.2 Global Water Cycle 2.01.2.1 Existence of Water on Earth 2.01.2.2 Water Cycle on the Earth 2.01.3 Global Water-Balance Requirements 2.01.3.1 Water Balance at Land Surface 2.01.3.2 Water Balance in the Atmosphere 2.01.3.3 Combined Atmosphere–River Basin Water Balance 2.01.3.3.1 Estimation of large-scale evapotranspiration 2.01.3.3.2 Estimation of total water storage in a river basin 2.01.3.3.3 Estimation of zonally averaged net transport of freshwater 2.01.3.4 Bottom Line of Global Water Balance 2.01.4 Global Water Balance 2.01.4.1 Uncertainties in Global Water-Balance Estimates 2.01.4.2 Water Balance and Climate 2.01.4.3 Annual Water Balance in Climatic Regions 2.01.5 Challenges in the Global Hydrology and Research Gaps 2.01.5.1 Macroscale Hydrological Modeling 2.01.5.2 Global Changes and Global Hydrology 2.01.5.3 Global Trace of Water Cycles 2.01.5.4 Interactions of Global- and Local-Scale Hydrology 2.01.5.5 Research Opportunities in Global Hydrology 2.01.5.6 Research Gaps in Global Hydrology Acknowledgments References

2.01.1 Introduction 2.01.1.1 The Earth System and Water The Earth system is unique in that water exists in all three phases, that is, water vapor, liquid water, and solid ice, when compared with the forms of water on other planets. The transport of water vapor is regarded as energy transport because of the large amount of latent-heat exchange that occurs during its phase change to liquid water (approximately 2.5 106 J kg1); therefore, the water cycle is closely linked to the energy cycle. Even though the energy cycle on the Earth is an open system driven by solar radiation, the amount of water on the Earth does not change during shorter than geological timescales (Oki, 1999), and the water cycle itself is a closed system. On a global scale, hydrologic cycles are associated with atmospheric circulation, which is driven by the unequal heating of the Earth’s surface and atmosphere in latitude (Peixo¨to and Oort, 1992). Annual mean absorbed solar energy at the top of the atmosphere is highest near the equator with approximately 300 W m2, and decreases rapidly at higher latitudes, and is

3 3 3 4 4 5 8 9 9 11 11 11 12 12 12 12 14 15 16 16 19 19 20 22 22 24 24

approximately 60 W m2 at the Arctic and Antarctic regions. Emitted terrestrial radiative energy from the Earth at the top of the atmosphere is approximately 250 W m2 for the areas between 201 N and 201 S, gradually decreases at higher latitudes, and is approximately 175 W m2 at the Arctic region and 150 W m2 at the Antarctic region. As a consequence, the net annual energy balance is positive (absorbing) over tropical and subtropical regions between 301 N and 301 S, and negative in higher latitudes (Dingman, 2002). Without atmospheric and oceanic circulations on the Earth, temperature differences on the Earth would have been more drastic. Temperatures in the equatorial zone would have been much higher such that the outgoing terrestrial radiation balances the absorbed solar energy, and the temperatures in the polar regions would have been much lower as well. Both the atmosphere and the ocean carry much energy from the equatorial regions toward the polar regions. In the case of atmosphere, the energy transport consists of sensible heat and latent-heat fluxes (Masuda, 1988). The global water circulation includes the latent-heat transport in which water vapor plays an active role in the atmospheric circulation. Water vapor is not a passive component of the atmosphere system;

3

4

Global Hydrology

rather, it affects atmospheric circulation by both radiative transfer and latent-heat release of phase change.

2.01.1.2 Water Reserves, Fluxes, and Residence Time The total volume of water on the Earth is estimated as approximately 1.4 1018 m3, and it corresponds to a mass of 1.4 1021 kg (Figure 1, revised from Oki and Kanae (2006)). Compared with the total mass of the Earth (5.974 1024 kg), the mass of water constitutes only 0.02% of the planet, but it is critical for the survival of life on the Earth, and the Earth is called the Blue Planet and the Living Planet. There are various forms of water on the Earth’s surface. Approximately 70% of its surface is covered with salty water, the oceans. Some of the remaining areas (continents) are covered by freshwater (lakes and rivers), solid water (ice and snow), and vegetation (which implies the existence of water). Even though the water content of the atmosphere is comparatively small (approximately 0.3% by mass and 0.5% by volume of the atmosphere), approximately 60% of the area of the Earth is always covered by cloud (Rossow et al., 1993). The Earth’s surface is dominated by the various phases of water. Water on the Earth is stored in various reserves, and various water flows transport water from one to another. Water flow

Water vapor over sea 10

Evaporation over ocean 436.5

(mass or volume) per unit time is also called water flux. The mean residence time in each reserve can be simply estimated from total storage volume in the reserve and the mean flux rate to and from the reserve:

Tm ¼ V=F

ð1Þ

where Tm, V, and F are mean residence time, total storage, and the mean flux rate, respectively. We can also represent the distribution of flux rate of water flow that comes in and goes out from the storage Chapman, 1972). The last column of Table 1 (simplified from the table in Korzun (1978)) presents global values of the mean residence time of water. Evidently, the water cycle on the Earth is a stiff differential system with variability on many timescales, from a few weeks to thousands of years. The mean residence time is also important when considering water-quality deterioration and restoration, since it can be an index of how much water is turned over. Apparently, river water or surface water is more vulnerable to pollution than groundwater; however, any measure to increase waterquality recovery tends to be more efficient for river water than groundwater, and, as suggested from Table 1, the mean

Total terrestrial precipitation 111 Snowfall Rainfall 12.5 98.5

Net water-vapor flux transport 45.5

Water vapor over land 3

Glaciers and snow 24 064

Total terrestrial evapotranspiration 65.5 21 Precipitation over ocean 391

6.4 11.7 Others (29.3)

7.6 11.6 Cropland (12.6) 2.66

Unirrigated

0.38 Domestic Sea 1 338 000

0.77

Irrigated

45.5

River 2

Industry

54

Forest (40.1)

0.2 0.3 Wetland (0.2) Wetland 11

31

Grassland (48.9)

29

Biological Permafrost water 300 1 Surface runoff 15.3

1.3 2.4 Lake (2.7) Soil moisture 17

Subsurface runoff 30.2

Lake 176

Groundwater 23 400 Flux, 103 km3 yr−1 Storage, 103 km3

The terrestrial water balance does not include Antarctica

( )

Area 106 km2

Figure 1 Global hydrological fluxes (1000 km3 yr1) and storages (1000 km3) with natural and anthropogenic cycles are synthesized from various sources. Big vertical arrows show total annual precipitation and evapotranspiration over land and ocean (1000 km3 yr1), which include annual precipitation and evapotranspiration in major landscapes (1000 km3 yr1) presented by small vertical arrows; parentheses indicate area (million km2). The direct groundwater discharge, which is estimated to be about 10% of the total river discharge globally, is included in river discharge. The values of area sizes for cropland and others are corrected from original ones. From Oki T, Nishimura T, and Dirmeyer P (1999) Assessment of annual runoff from land surface models using total runoff integrating pathways (TRIP). Journal of the Meteorological Society of Japan 77: 235–255 and Ok T and Kanae S (2006) Global hydrological cycles and world water resources. Science 313(5790): 1068–1072.

Global Hydrology Table 1

5

World water reservesa

Form of water

Covering area (km2)

Total volume (km3)

Mean depth (m)

World ocean Glaciers and permanent snow cover Ground water Ground ice in zones of permafrost strata Water in lakes Soil moisture Atmospheric water Marsh water Water in rivers Biological water Artificial reservoirs Total water reserves

361 300 000 16 227 500 134 800 000 21 000 000 2 058 700 82 000 000 510 000 000 2 682 600 148 800 000 510 000 000

1 338 000 000 24 064 100 23 400 000 300 000 176 400 16 500 12 900 11 470 2120 1120 8000 1 385 984 610

3700 1463 174 14 85.7 0.2 0.025 4.28 0.014 0.002

510 000 000

2718

Share (%) 96.539 1.736 1.688 0.0216 0.0127 0.0012 0.0009 0.0008 0.0002 0.0001

Mean residence time 2500 years 1600 years 1400 years 10 000 years 17 years 1 year 8 days 5 years 16 days A few hours 72 days

100.00

a

Simplified from Table 9 of ‘‘World water balance and water resources of the earth’’ by UNESCO Korzun, 1978. The last column, mean residence time, is from Table 34 of the report.

residence time of river water is shorter than that of groundwater. Since the major interests of hydrologists have been the assessment of volume, inflow, outflow, and the chemical and isotopic composition of water, the estimation of mean residence time of a certain domain has been one of the major targets of hydrology. It should be recalled that the residence time estimated with isotope tracers often differs from the hydrological residence time derived from Equation (1) (Uhlenbrook et al., 2002, 2004). This is due to the fact that in the subsurface system, the diffusive exchange processes between mobile and immobile parts make the residence time usually much longer. This process is particularly important for hydrochemical processes.

2.01.2 Global Water Cycle 2.01.2.1 Existence of Water on Earth Table 1 and Figure 1 denote the quantity of water stored in each of the reserves on the Earth. Most of the storage values given in Table 1 are taken from Korzun (1978), except for water vapor in the atmosphere which is calculated from atmospheric data (Oki et al., 1995). The various reserves of water on the Earth are discussed in the following: The proportion of water in the ocean is large (96.5%). Even though in classical hydrology, ocean processes are traditionally excluded, the global hydrological cycle is never closed without including them. The ocean circulations carry large amounts of energy and water. The surface ocean currents are driven by surface wind stress, and the atmosphere itself is sensitive to the sea-surface temperature. Temperature and salinity determine the density of ocean water, and both factors contribute to the overturning and deep-ocean general circulation Other major reserves are solid waters on the continent (glaciers and permanent snow cover) and groundwater. Glaciers are the accumulation of ice originated from the atmosphere, and they generally move slowly on land over a long period. Glaciers form a discriminative U-shaped valley over land, and remain moraine when they retreat. If a glacier flows into an ocean, its terminated end often forms an iceberg. Glaciers react in comparatively longer timescales against

climatic change, and they also induce isostatic responses of continental-scale upheavals or subsidence in even longer timescales. Even though it is predicted that the thermal expansion of oceanic water dominates the anticipated sea-level rise due to global warming, glaciers over land are also a major concern as the cause for sea-level rise associated with global warming. Groundwater is the subsurface water occupying the saturated zone. It contributes to runoff in the low-flow regime between storm events. Deep groundwater may also reflect the long-term climatological situation. Groundwater in Table 1 includes both gravitational and capillary water, but groundwater in the Antarctica (roughly estimated as 2 106 km3) is excluded. Gravitational water is the water in the unsaturated zone (vadose zone) which moves under the influence of gravity. Capillary water is water found in the soil above the water table by capillary diffusion, a phenomenon associated with the surface tension of water in soils acting as porous media. In terms of groundwater recharge, Do¨ll and Fiedler (2008) estimated the global groundwater recharge flux to be 12 666 km3 yr1 and approximately 1000 mm yr1 in the Amazon region. They assumed the recharge flux is a fraction of the total runoff. Koirala (2010) estimated groundwater recharge flux by coupling a land-surface scheme, namely Minimal Advanced Treatments of Surface Interaction and RunOff (MATSIRO; see Takata et al., 2003), with a macro-scale groundwater representation Yeh et al., 2005). The global distribution of model-simulated groundwater recharge is illustrated in Figure 2(a). Total groundwater recharge flux is estimated as 31 789 km3 yr1 and the value is close to the flux of subsurface runoff in Figure 1 (30 200 km3 yr1). Soil moisture is the water that is held above the groundwater table. It influences the energy balance at the land surface, such that a lack of available moisture suppresses evapotranspiration (which consists of soil evaporation, plant transpiration, and interception loss), and changes surface albedo. Soil moisture also alters the fraction of precipitation partitioned into direct runoff and infiltration. The precipitation water becoming direct runoff cannot be evaporated from the same place, while the water infiltrated into soil may be taken up by hydraulic suction and evaporated back into the atmosphere. The global distribution of model-estimated mean

6

Global Hydrology 90° N

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Figure 2 (a) Global map of long-term mean net groundwater recharge (mm yr ) estimated by an LSM coupled with macroscale groundwater model (Yeh and Eltahir, 2005; S. Koirala, 2010); (b) global distribution of annual mean soil wetness index estimated by 13 LSMs averaged for 1986–95 through the second phase of the Global Soil Wetness Project; (c) same as (b) but for annual precipitation (mm yr1) used in the GSWP2 based on observation; (d) same as (b) but for estimated annual runoff (mm yr1) by LSMs; (e) same as (b) but for mean river discharge (106 m3 yr1); (f) annual vapor-flux convergence (mm year1) for 1989–92 (Oki et al., 1995). (g) Annual mean evapotranspiration (mm yr1) estimated as a residual of (f) and precipitation corresponding to the period; (h) same as (b) but for annual mean evapotranspiration (mm yr1) by LSMs for 1986–95. Data of (a) from Takata K, Emori S, and Watanabe T (2003) Development of minimal advanced treatments of surface interaction and runoff. Global and Planetary Change 38: 209–222 and Koirala S (2010) Explicit Representation of Groundwater Process in a Global-Scale Land Surface Model to Improve Hydrological Predictions. PhD Thesis, The University of Tokyo; (b) from Dirmeyer PA, Gao XA, Zhao M, Guo ZC, Oki T, and Hanasaki N (2006) GSWP-2 multimodel anlysis and implications for our perception of the land surface. Bulletin of the American Meteorological Society 87: 1381–1397; and (f) from Oki T, Musiake K, Matsuyama H, and Masuda K (1995) Global atmospheric water balance and runoff from large river basins. Hydrological Processes 9: 655–678.

Global Hydrology

7

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Figure 2 Continued.

soil wetness index is shown in Figure 2(b). Generally, the distribution is correlated with precipitation distribution (Figure 2(c)) as well as with runoff distribution (Figure 2(d)), but the global distribution of river discharge (Figure 2(e)) accumulates total runoff generated in the upper watershed, and the shape of the river channels can be seen in the distribution. The atmosphere carries water vapor, which influences the heat budget via latent-heat-exchange processes. Condensation of water vapor releases latent heat, which warms the atmosphere and affects the atmospheric general circulation. Liquid water (droplets, clouds, etc.) in the atmosphere is another result of condensation. Clouds significantly change the radiation in the atmosphere and at the Earth’s surface. However, the volume of liquid (and solid) water contained in the atmosphere is relatively small, as most of the water in the atmosphere exists as water vapor. Water vapor is also the major absorber in the atmosphere of both short-wave and long-wave radiation. Precipitable water is the total water vapor in the atmospheric column integrated from land surface to the top of the atmosphere. Vertically integrated water-vapor flux convergence is a useful tool to diagnose global water balance (see Figure 2(f) for its global distribution). The amount of water stored in rivers (Figure 2(e)) is rather tiny compared to other reserves at any time; however, the recycling speed, which can be estimated as the inverse of the mean residence time (Equation (1)), of river water (river discharge) is relatively high, and it is important because most societal applications ultimately depend on river water as a renewable and sustainable resource. The amount of water stored transiently in a soil layer, in the atmosphere, and in the river channels is relatively minute,

and the time spent through these subsystems is relatively short. However, they play a dominant role in the global hydrological cycle.

2.01.2.2 Water Cycle on the Earth The water cycle plays many important roles in the climate system, and Figure 1 schematically illustrates various flow paths of water in the global hydrologic system (Oki and Kanae, 2006). Precipitation is calculated from global estimates based on observations from the forcing data of the Global Soil Wetness Project (GSWP2, the second phase the project, see the discussion in Section 2.01.4) over land, and data from Climate Modeling Analysis and Prediction (CMAP; Xie and Arkin, 1996) over ocean. Land-surface fluxes, such as evapotranspiration and surface and subsurface runoff, are the estimated results from GSWP2. Differentiation of precipitation between snow and rain over land is also either estimated by land-surface models that participated in the GSWP2, or given by individual forcing determined by temperature. Values on human water withdrawals for irrigation, industry, and households are taken from Shiklomanov (1997). Water-vapor transport and its convergence are estimated using the European Centre for Medium-Range Weather Forecast (ECMWF) objective analyses, obtained as the 4-year mean from 1989 to 1992 (Oki et al., 1995). The roles of these water fluxes in the global hydrologic system are now briefly reviewed:

•

Precipitation is the water flux from atmosphere to land or ocean surface. It drives the hydrological cycle over the land surface, also changes surface salinity (and temperature) over the ocean, and affects its thermohaline circulation. Rainfall

8

Global Hydrology

Annual runoff 210°

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Water scarcity index

Figure 2 Continued.

•

refers to the liquid phase of precipitation. A part of it is intercepted by canopy over vegetated areas, and the remaining part reaches the Earth‘s surface as through-fall. The highly variable, intermittent, and concentrated behavior of precipitation in time and space domain compared to other major hydrological fluxes mentioned below makes the observation of this quantity and the aggregation of the process complex and difficult. Global distribution of precipitation is presented in Figure 2(c). Currently, satellitebased estimates merged with in situ observational data have been produced and revealed to the public (e.g., Kubota et al., 2007). Snow has special characteristics compared with rainfall. Snow may be accumulated and the surface temperature will not rise above 0 1C until the completion of snowmelt. The albedo of snow is quite high (as high as clouds). Consequently, the existence of snow changes the surface energy and water budget enormously. A snow surface typically

•

reduces the aerodynamic roughness, and therefore may also have a dynamical effect on the atmospheric circulation and hydrologic cycle. Evaporation is the return flow of water from the surface to the atmosphere and the latent-heat flux from the surface. The amount of evaporation is determined by both atmospheric and hydrological conditions. From the atmospheric point of view, the partition of incoming solar energy to the surface between latent and sensible heat flux is important. Wetness at the surface influences this partition significantly because the ratio of actual evapotranspiration to the potential evaporation is reduced due to drying stress. The stress is sometimes formulated as a resistance under which evaporation is classified as hydrology driven (soil controlled). If the land surface is wet enough compared to available energy for evaporation, the condition is classified as radiation driven (atmosphere controlled).

Global Hydrology

9

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•

•

Transpiration is the evaporation of water through stomata of leaves. It has two special characteristics different from evaporation from soil surfaces. One is that the resistance of stomata is related not only to the dryness of soil moisture but also to the physiological conditions of vegetation, through the opening and closing of stomata. Another feature is that roots of plants can uptake water from deeper soil layers and transpirate the water, compared to the case of evaporation from bare soil without plants. Vegetation also modifies the surface energy and water balance by altering surface albedo and by intercepting some precipitation and evaporating this rainwater. The global distribution of total evapotranspiration is shown in Figure 2(g), which is estimated using the atmospheric water-balance computation (Equation (7) in Section 2.01.3.3.1), and in Figure 2(h), estimated by land-surface models (in Section 2.01.4.1). Runoff returns water from the land to the ocean, which may otherwise be transported in vapor phase by evaporation and atmospheric advection. The runoff into the ocean is also important for the freshwater balance and the salinity of the ocean. Rivers carry not only water mass but also sediment, chemicals, and various nutritional matters from continents to seas. Without rivers, global hydrologic cycles on the Earth are not closed. Runoff at hillslope scale is a nonlinear and complex process. Surface runoff can be generated when the intensity of rainfall or snowmelt exceeds the infiltration rate of the soil, or when precipitation falls on the saturated land surface. Saturation at land surface can be formed mostly by topographic-concentration mechanism along the hillslopes. Infiltrated water in the upper part of the hillslope flows down the slope and discharges at the bottom of the hillslope. Due to the highly

variable heterogeneity of topography, soil properties (such as hydraulic conductivity and porosity) and precipitation, basic equations such as Richard’s equation, which are fairly valid at a point scale or hillslope scale, cannot be directly applied in the macroscale because of the nonlinearity involved. The global distributions of runoff and river discharge are illustrated in Figures 2(d) and 2(e). The global water cycle integrates these components, which consist of the state variables (precipitable water, soil moisture, etc.) and the fluxes (precipitation, evaporation, etc.).

2.01.3 Global Water-Balance Requirements The conservation law of water mass in any arbitrary control volume indicates the water balance. In this section, the water balance over land, for an atmospheric column, and its combination is presented (Oki, 1999). Some applications of these water-balance equations for estimating some of the waterbalance components are introduced as well.

2.01.3.1 Water Balance at Land Surface In the field of hydrology, river basins have commonly been selected for study, and water balance has been estimated using ground observations, such as precipitation, runoff, and storage in lakes and/or groundwater. The water balance over the land is described as

q S=q t ¼ P  E  Ro  Ru

ð2Þ

where S represents the water storage within the area, t is the time, (q S/q t) is the change of total water storage with time,

10

Global Hydrology Annual evapotranspiration (mm yr−1) 0

60° E

120° E

180°

120° W

60° W

0

90° N

60° N

30° N

EQ.

30° S

60° S

90° S

(g)

0.00

300

900

600

1200

1500

1800

2100

2400

2700

3000

90°

60°

30°

0°

−30°

−60°

−90° −180°

(h)

−200

−150°

0

−120°

20

−90°

50

−60

100

−30°

150

0°

200

30°

300

500

60°

800

90°

1400

120°

1800

150°

180°

2400

3000

Figure 2 Continued.

P is the precipitation, E is the evapotranspiration, Ro is the surface runoff, and Ru is the groundwater movement (all fluxes above are given in the unit volume per time step). S includes snow accumulation in addition to soil moisture, groundwater, and surface-water storage including retention

water within the control volume. The control volume is defined by the area of interest over the land with its bottom generally at the impermeable bedrock. These terms are shown in Figure 3(a). Equation (2) implies that water storage over land is increased by precipitation, and decreased

Global Hydrology

by evapotranspiration, surface runoff, and groundwater movement. If the considered area of water balance is set within an arbitrary boundary, Ro represents the net outflow of water from this area (i.e., the total outflow minus total inflow from surrounding areas). Although, in general, it is not easy to estimate groundwater movement Ru, the net flux per unit area within a large area is expected to be comparatively small. If all groundwater movement is considered to be that observed at the gauging point of a river (Ru ¼ 0), then Equation (2) becomes

q S=q t ¼ P  E  Ro

2.01.3.2 Water Balance in the Atmosphere Atmospheric water-vapor flux convergence contains waterbalance information that can complement the traditional hydrological elements such as precipitation, evapotranspiration, and discharge. The basic concepts as well as the application of atmospheric data to estimate terrestrial water balance were first presented by Starr and Peixo¨to (1958). The atmospheric water balance for a column of atmosphere from the bottom at land surface to the top of the atmosphere is described by

q W=q t ¼ Q þ ðE  PÞ

2.01.3.3 Combined Atmosphere–River Basin Water Balance Since there are common terms in Equations (3) and (4), they can be combined as

q W=q t þ Q ¼ ðP  EÞ ¼ q S=q t þ Ro

• •

Annual change of atmospheric water-vapor storage is negligible ((q W/q t) ¼ 0). Annual change of water storage at the land is negligible ((q S/q t) ¼ 0).

With these assumptions, Equation (5) simplifies into

Q ¼ ðP  EÞ ¼ Ro

(a) Water balance

in the basin

Groundwater movement

ð6Þ

If a river basin is considered as the water-balance region, Ro is simply the discharge from the basin. The simplified Equation (6) demands that the water-vapor convergence, that is, precipitation minus evaporation and net runoff should balance over the annual period when the temporal change of all storage terms can be neglected.

Precipitable water

Vapor flux

Precipitable water

Runoff Basin storage

ð5Þ

Figure 3(c) illustrates the balance in Equation (5), and shows that the difference of precipitation and evapotranspiration is equal to the sum of the decrease of atmospheric water-vapor storage and lateral (horizontal) convergence, and also to the sum of the increase of water storage over the land and runoff. Theoretically, Equation (5) can be applied for any control volume of land area combined with the atmosphere above, even though the practical applicability depends on the accuracy and availability of atmospheric and hydrologic information. The following further assumptions are often employed in annual water-balance computations:

ð4Þ

where W represents the precipitable water (i.e., column storage of water vapor) and Q is the convergence of water-vapor flux in the atmosphere (all fluxes given in the unit volume per time step). Since the atmospheric water content in both solid and liquid phases are generally small, only the water vapor is considered in Equation (4). The balance is schematically illustrated in Figure 3(b), which describes that the water storage

Precipitation Evapotranspiration

in an atmospheric column is increased by the lateral convergence of water vapor and evapotranspiration through the bottom of the column (i.e., land surface), and decreases by the precipitation falling out from the bottom of the atmosphere column to the land.

ð3Þ

This assumption is generally valid at the outlet of a catchment. In most cases, surface runoff Ro becomes river discharge through the transport of river-channel network. The river discharge is an integrated quantity over the whole catchment and can be observed at a downstream point in contrast to other fluxes, such as P and E, which have to be spatially measured.

11

Vapor flux

Runoff

Precipitation

Basin storage

Evapotranspiration (b) Water balance in the atmosphere

Groundwater movement

(c) Combined water balance

Figure 3 (a) Terrestrial water balance, (b) atmospheric water balance, and (c) combined atmosphere–land surface water balance. (a), (b), and (c) correspond to Equations (2), (4), and (5), respectively.

12

Global Hydrology

2.01.3.3.1 Estimation of large-scale evapotranspiration Generally, it is not an easy task to obtain large-scale evapotranspiration E based on observations except for the annual timescale in which E can be estimated as the residual of P and Ro. However, the combined water balance can help estimate E at a shorter timescale, for example, monthly. Note that Equation (4) can be rewritten as

E ¼ q W=q t  Q þ P

ð7Þ

which can be applicable over a period shorter than a year, unlike the assumption in Equation (6). If atmospheric and precipitation data are available over a short timescale such as a month or a day, evapotranspiration can be estimated at the corresponding timescales; however, it is also subject to severe limitations imposed by the data accuracy. The region over which the evapotranspiration is estimated is not limited to a river basin; rather, it depends on the scale and the associated accuracy of the available atmospheric and precipitation data. The global distributions of precipitation P, integrated water-vapor convergence Q, and evapotranspiration E estimated by Equation (7) are presented in Figures 2(c), 2(f), and 2(g), respectively. The zonal mean precipitation P, integrated water-vapor convergence Q, and evapotranspiration E estimated by Equation (7) are presented in Figures 4(a)–4(c). As can be seen from the zonal mean precipitation along the midlatitude, storm tracks over the North Pacific and Atlantic oceans are stronger in December–January–February (DJF) than in June–July–August (JJA). The Intertropical Convergence Zone (ITCZ) is enhanced in JJA, when the southeastern part of the Asian continent is covered by the southwest monsoon rainfall. The distribution of E is less dependent on the latitude and has smaller seasonal changes compared to P (see also Trenberth and Guillemot, 1998). The mean evapotranspiration in tropical areas is approximately 4 mm d1.

2.01.3.3.2 Estimation of total water storage in a river basin

2.01.3.4 Bottom Line of Global Water Balance Water balance on the global scale with consideration of land and ocean areas separately can be expressed as

Total terrestrial water storage, as the sum of surface water (such as river water, snow water, and water in lakes), soil moisture, and groundwater, is generally difficult to estimate on the global scale. Combining Equations (3) and (4) yields

q S=q t ¼ q W=q t þ Q  Ro

the cycles of energy and water are closely related. Wijffels et al. (1992) used values of atmospheric water-vapor convergence Q from Bryan and Oort (1984) and discharge data from Baumgartner and Reichel (1975) to estimate the freshwater transport by oceans and atmosphere, but their results seem to have large uncertainties and they did not present the freshwater transport by rivers. The annual freshwater transport in the meridional (north– south) direction can be estimated from Q and river discharge with geographical information such as the location of river mouths and basin boundaries (Oki et al., 1995). The estimated result is shown in Figure 5, in which it shows that in the case of oceans net transport is the residual of northward and southward freshwater flux by all ocean currents globally, and it cannot be compared directly with individual ocean currents such as the Kuroshio and the Gulf Stream. Transports by the atmosphere and by the ocean have almost the same absolute values for most of the latitudes, but with a different sign. The transport by rivers is about 10% of these other fluxes globally (there may be an underestimation because average Q tends to be smaller than average river discharge observed at the global land surface). The negative (southward) peak by rivers at 301 S is mainly due to the Parana River in South America, and the peaks at the equator and 101 N are due to rivers in South America, such as the Magdalena and Orinoco. Large Russian rivers, such as the Ob, Yenisey, and Lena, carry freshwater toward the north between 50 and 701 N. These results indicate that the hydrological processes over land play non-negligible roles in the climate system, not only by the exchange of energy and water at the land surface, but also through the transport of freshwater by rivers which affects water balance of the oceans and forms a part of the hydrological circulation on the Earth among the atmosphere, land, and oceans.

ð8Þ

which indicates that the change of water storage in the control volume over the land can in principle be estimated from the atmospheric and runoff data. Although an initial value is required to obtain the absolute value of storage, the atmospheric water balance can be useful in estimating the seasonal change of total water storages in large river basins.

2.01.3.3.3 Estimation of zonally averaged net transport of freshwater The meridional (north–south direction) distribution of the zonally averaged annual energy transports by the atmosphere and the ocean has been evaluated, even though there are quantitative problems in estimating such values (Trenberth and Solomon, 1994). However, the corresponding distribution of water transport has not often been studied, although

P1  E1  q S1 =q t ¼ R ¼ ðPo  Eo Þ þ q So =q t

ð9Þ

where P, E, S, and R represent precipitation, evapotranspiration, total water storage, and continental runoff, respectively, with the subscript l indicating values for land and o for ocean. For the steady state, the temporal changes of Sl and So can be neglected and Equation (9) becomes

P1  E1 ¼ R ¼ ðPo  Eo Þ

ð10Þ

which indicates that continental runoff can be estimated as a residual of total evapotranspiration (Eo) and precipitation (Po) over the ocean. It could be an effective method to estimate continental runoff since a macroscale estimation of precipitation and evaporation is relatively easier over ocean than over the land. If precise estimates of the long-term trend of global mean precipitation and evapotranspiration over both land and ocean are available, there is a potential to infer the trend of the water stored over the land or ocean as suggested by Equation (9).

Global Hydrology

13

2.01.4 Global Water Balance

From Equation (10)

Pe ¼ Ee

ð11Þ

can be derived which states that precipitation all over the Earth Pe ¼ Pl þ Po and evapotranspiration all over the Earth Ee ¼ Pl þ Po should be identical under the conditions when the temporal changes of water storage over land and ocean are negligible.

The values quoted in Table 1 and Figure 1 are estimated based on various observations with some assumptions in order to obtain global perspectives. These values are sometimes different in other references probably because the source of observed data, methodology of estimation, and assumptions are different. In some cases, global water balances are estimated using empirical relationship of evapotranspiration to precipitation in each latitude belt (Baumgartner and Reichel, 1975).

Zonal mean precipitation (overall, land and sea) 10.0 Annual DJF JJA

(mm d−1)

7.5

5

2.5

0 60° S (a)

40° S

20° S

90° S

EQ.

20° N

40° N

60° N

Latitude

90° N

Vapor-flux convergence (overall, land and sea) 6 Annual DJF JJA

(mm d−1)

3

0

−3

−6 (b)

60° S 90° S

40° S

20° S

EQ. Latitude

20° N

40° N

60° N 90° N

Figure 4 (a) Meridional distribution of precipitation (P ) for mean over land and sea; (b) same as (a) but for vapor-flux convergence (Q ); (c) same as (a) but for evapotranspiration (E ) calculated as a residual of P and Q. Data of (a) from Xie P and Arkin PA (1996) Analyses of global monthly precipitation using gauge observations, satellite estimates, and numerical model predictions. Journal of Climate 9: 840–858 and (b) from Oki T, Musiake K, Matsuyama H, and Masuda K (1995) Global atmospheric water balance and runoff from large river basins. Hydrological Processes 9: 655–678.

14

Global Hydrology Zonal mean evaporation (overall, land and sea) 9 Annual DJF JJA

(mm d−1)

6

3

0

−3 (c)

60° S 90° S

40° S

20° S

EQ. Latitude

20° N

40° N

60° N 90° N

Figure 4 Continued.

Under an international research project, the land-surface models (LSMs) were used to estimate global water and energy balances for 1986–95 in order to obtain global distribution of surface soil moisture, which is not easy to obtain but relevant for understanding the land–atmosphere interactions (Dirmeyer et al., 2006). The project was called the second phase of GSWP2 and its goal was to produce state-of-the-art global data sets of land-surface fluxes, state variables, and related hydrologic quantities.

2.01.4.1 Uncertainties in Global Water-Balance Estimates In GSWP2, meteorological forcing data are hybrid products of the National Center for Environmental Prediction (NCEP)/ Department of Energy (DOE) reanalysis data and observational data based on in-situ and satellite monitoring, provided at a 3-hourly time step for a period of 13.5 years from July 1982 to December 1995. The first 3.5 years’ data are used for spin up. The land-surface parameters are specified from the Earth Resources Observation and Science Data Center (EDC) for the land-cover data and the International Geosphere–Biosphere Programme Data Information System (IGBP-DIS) for the soil data. Both land-surface parameters and meteorological forcing are at 11 resolution for all land grids excluding Antarctica. Figure 6(a) illustrates the model-derived global water balance over the global land excluding ice, glacier, and lake. The numeric in the box corresponds to the 10-year mean annual value of eight LSMs participated in the GSWP2 project (Oki et al., 2005). All the simulations were performed using identical forcing data given by 11 11 longitudinal and latitudinal grid boxes, and typical time steps of the calculations are 5 min to 3 h. The vertical ranges shown above and below the boxes indicate the maximum and minimum values in the interannual variation of mean annual value among the eight LSMs. The horizontal ranges shown left and right to the

boxes indicate the maximum and minimum values of the intermodel variation of the 10-year mean value of eight LSMs. Generally, intermodel variation exceeds interannual variation, which suggests that the uncertainty associated with model selection is larger than the sampling error of estimating global water balance. In the case of rainfall, intermodel variation is small because identical precipitation forcing was given to LSMs so that the differences among LSM estimates were merely caused by the rain/snow judgment made by each modeling group. The advantage of using models to estimate global water balance is the capability to have more detailed insights than using observations. For example, snow over the land excluding ice and glacier areas is approximately 10% of total precipitation, and the ratio of surface runoff and subsurface runoff is approximately 2:3 in Figure 6(a), but the latter is approximately 1:2 in Figure 1. This is because the model results used for calculating average value are different for Figures 1 and 6(a). In some LSMs, neither surface nor subsurface runoff process is considered, which is the reason why the minimum values are zero. Even though at present it is difficult to assess the validity of these breakdowns, due to the lack of observations on the partitions between snow and rain or between surface and subsurface runoff, such model-based estimates will stimulate scientific interest to collect and compile global information on these important hydrological quantities in the future. Further, evapotranspiration was estimated separately by bare-soil evaporation (Es), evaporation from intercepted water on leaves (Ei), evaporation from open water (Ew), and transpiration from vegetation (Et), as shown in Figure 6(b), even though the intermodel variations are quite large partially because some LSMs do not consider all of these components of evapotranspiration. Even though the values in Figure 6(b) are not definitive, it is interesting to see that bare-soil evaporation

Global Hydrology

15

Toward north (mean 1989−92) 40

100

Atmosphere Continent

30

Ocean

10

0

50

Land/sea (%)

Water flux (1012 m3 yr−1)

20

−10 −20 −30 −40

0 60° S 90° S

30° S

EQ. Latitude

30° N

60° N 90° N

Figure 5 The annual freshwater transport in the meridional (north–south) direction by atmosphere, ocean, and rivers (land). Water-vapor flux transport of 20  1012 m3 yr1 corresponds to approximately 1.6  1015 W of latent-heat transport. Shaded bars behind the lines indicate the fraction of land at each latitudinal belt.

and transpiration from vegetation are closely comparative, and interception loss is approximately 10% of the total evapotranspiration. It would be interesting if these estimates can be revised and validated by certain observation-based measures, and intermodel discrepancies can therefore be reduced.

2.01.4.2 Water Balance and Climate Based on observed precipitation (Xie and Arkin, 1996) and river-discharge records archived at the Global Runoff Data Centre, annual water balance for most river basins worldwide was estimated (Oki et al., 1999), and this is presented in Figure 7. The ordinate in Figure 7 is the residual of long-term mean annual precipitation and runoff, and this annual loss should correspond to long-term mean annual evapotranspiration. Different symbols are used for plotting: red stars indicate water balance of the river basins where gauging stations are located between 201 S and 201 N. The plus symbols indicate the river basins where gauging stations are located between 201 and 401 in both hemispheres, and the blue circles are 401 or higher. The line connects the mean precipitation and annual loss for each 51 latitudinal belt. As seen, even though the scatter is large, approximately 70% of precipitation is evapotranspirated in high-latitude river basins. On the other hand, mean evapotranspiration in tropical river basins is approximately 1000 mm yr1 with less dependency on annual precipitation. Such analyses on the relationship between P and E have long been used for estimation of global water balance, for example, in Baumgartner and Reichel (1975). It is also

clear from Figure 7 that river basins with annual precipitation less than 800 mm yr1 have marginal amounts of river runoff since most precipitation is used for evapotranspiration. In these river basins, evapotranspiration is mainly controlled by the availability of water (water controlled), and this is in contrast to tropical river basins with precipitation higher than 1000 mm yr1 where annual evapotranspiration is limited by the available energy (energy- or radiation controlled). Budyko (1974) proposed an equation

E=P ¼ ½zðtanh 1=zÞð1  cosh z þ sinh zÞ ð1=2Þ

ð12Þ

where E and P are annual evapotranspiration and precipitation respectively, and the Budyko’s dryness index is defined as

z ¼ Rn =lP

ð13Þ

where Rn and l are net radiation and the coefficient of latent heat, respectively. This equation is derived by considering that the E/P should be asymptotic to 1.0 for dry regions (large z) since E should be less than P, and E/P should be asymptotic to Rn/lP for wet regions (small z) since E should be less than Rn/l. Budyko’s equation (12) is conceptual, but it can provide a realistic water balance as shown in Figure 8. Mean water balance averaged for each 0.2z( ¼ Rn/lP) bin estimated by an LSM corresponds fairly well with the curve according to the Budyko’s equation (12), even though large scatters are found in the plots of each 11 longitudinal and latitudinal grid box. Yang et al.

16

Global Hydrology

Global terrestrial water budget Unit: mm yr−1 91 58

86

793

102

80

Legend

759

742

Snow

508

414

726

742

574

499

Rainfall

0

Intermodel range

142

302

133

Surface runoff

432

338 308

Total discharge

138 60

303

150

Soil-water storage

149

* Except lake and ice/glacier, 14 409 grids or 1.302 59e + 8 km2

Average of eight models (1986−95)

364 270

152

Subsurface runoff

788

726

302

196

0

759

ET 151

214

Interannual range

793

524

788

Balance = −1 mm

(a)

Global composition of evapotranspiration Unit: mm yr−1

Legend

574

793

Interannual range

524 414

508 499

742

ET Es 0

65

62

137

0

229

544

Average of eight models (1986−95)

248 0 7

238

Intermodel range

Et

59

11

788

726

Ew

Ei

759

6

47

210

531

235

1

223

(b)

Figure 6 (a) Global terrestrial water balance averaged for 1986–95 estimated by eight LSMs in boxes. Interannual variation range (vertical) for 1986–95 and intermodel discrepancies (horizontal) among eight models are presented for the annual mean estimates; (b) same as (a) but for global composition of evapotranspiration.

(2009) analyzed annual water balance in 99 river basins in China and concluded that this scatter can at least partially be explained by the vegetation cover in the river basin.

2.01.4.3 Annual Water Balance in Climatic Regions Annual water balances estimated by GSWP2 were analyzed and each 11 11 grid box was classified into one of the

following six climatic regions according to the Budyko’s dryness index z and annual precipitation:

• • • • •

arid region: z44.0; semiarid region: 4.04z4 ¼ 2.0; semi-humid region: 2.04z4 ¼1.2; humid region: 1.24z4 ¼ 0.7; tropical humid region: 0.74 ¼ z and annual precipitation larger than 2000 mm yr1; and

Global Hydrology

17

Precipitation and annual loss over the world 2000 Low latitude (20°S−20°N) Mid-latitude (20°S−40°S, 20°N−40°N) High latitude (40°S−90°S, 40°N−90°N)

Annual loss (mm yr−1)

1600

1200 5 15 20 25

800

10

0

30 40 45 35 50 55 60 65

400 70 0 0

400

800

1200

1600

Precipitation (mm

2000

2400

2800

yr−1)

Figure 7 Annual water balance in major river basins. Annual loss is estimated as a residual of annual precipitation and observed runoff in catchments of 250 gauging stations of river discharge. From Oki T, Nishimura T, and Dirmeyer P (1999) Assessment of annual runoff from land surface models using total runoff integrating pathways (TRIP). Journal of the Meteorological Society of Japan 77: 235–255.

•

very humid region: 0.74 ¼ z and annual precipitation less than 2000 mm yr1.

The six classified regions are illustrated in Figure 9 along with the ice-covered region. The differentiation is difficult between tropical humid region and very humid region only by using the Budyko’s dryness index z. Therefore, annual precipitation is considered in the classification. It is interesting to see that z is similar in both tropical and high-latitude regions in addition to the Asian monsoon region. Perhaps, it is also necessary to consider the seasonal change of major waterbalance terms for better differentiation of these regions. Long-term mean water balance for each climatic region classified by z is presented in Figure 10(a). Separation of mean annual precipitation into evapotranspiration, surface runoff, and subsurface runoff is also illustrated. The sum of these corresponds to annual precipitation, and the ratio of annual evapotranspiration to precipitation is close to the mean z of each region. Slightly negative evapotranspiration in the ice region indicates net sublimation in the region, that is, the land surface obtains energy from the atmosphere through sublimation. Evapotranspiration is also divided into four components: bare-soil evaporation, transpiration, evaporation from intercepted water, and evaporation from open water, as presented in Figure 10(b). Note that not all the LSMs that participated in GSWP2 have considered all of these four components, and some of these values could be underestimated. Transpiration

and evaporation from intercepted water are proportional to the vegetation biomass in each region, and it is interesting to note that the magnitude of bare-soil evaporation is relatively uniform globally than other components, except for arid, very humid, and ice regions.

2.01.5 Challenges in the Global Hydrology and Research Gaps 2.01.5.1 Macroscale Hydrological Modeling The development of macroscale hydrological models was a serious topic of discussion among Japanese scientists researching land–atmosphere interaction studies in the early 1990s when Global Energy and Water Experiment (GEWEX) Asian Monsoon Experiment (GAME) was under preparation. Two approaches were identified: one to expand a conventional microscale rainfall-runoff hydrological model into a macroscale model, which can run on the continental scale with a detailed energy balance and vegetation representation, and the other to enhance hydrological processes in LSMs and couple them with horizontal water-flow processes, particularly with river flow. A river-routing scheme was hence developed with a globalflow direction map, and named as the total runoff integrating pathways (TRIPs) (Oki and Sud, 1998). Such a river-routing scheme can be coupled with any LSM, and can also be used as a post-processor integrating the runoff estimated by LSMs into

18

Global Hydrology Budyko’s diagram for GSFC (Mosaic) (1987)

E/P; evapotranspiration/precipitation

1.0

0.8

0.6

0.4

0.2

0.0 0.0

0.5

1.0

1.5 2.0 2.5 3.0 3.5 4.0 Rn/lP; net radiation/(latent heat * precipitation)

4.5

5.0

5.5

6.0

Figure 8 Annual energy and water balance of 11 longitude and latitudinal grids calculated by an LSM for 1987 (Koster et al., 1999). Plots indicate the energy and water balances in each grid box and green line presents the mean E/P for each 0.2 z ¼ Rn/lP bin. The red curve indicates Equation (12). From Budyko MI (1974) Climate and Life, Miller DH (trans.). San Diego, CA: Academic Press.

Arid Semiarid Semi-humid Humid Tropical humid Very humid Iee cover Figure 9 Classification of energy and water-balance regime using Budyko’s dryness index z ¼ Rn/lP.

Mean water balance in climatic regions (1986−95)

l ba

e

G lo

Ic

id m

id Ve

ry

hu

um

id

lh

Tr o

pi

Se

ca

m

i- h

H um

r id m

ia

Ar Se

(a)

id

Rsub Rsurf ET

um

3000 2500 2000 1500 1000 500 0 −500

id

(mm y−1)

Global Hydrology

1200 Ew Ei Et Es

1000 800 600

19

seasonal pattern of observed TWS variation by Gravity Recovery and Climate Experiment (GRACE; see Tapley et al., 2004) without an appropriate representation of river-storage component. The dominant role of river storage was already indicated in a pilot study which compared total TWS changes estimated by the atmospheric water-balance method and a GCM simulation coupled with TRIP in the Amazon river basin (Oki et al., 1996). However, the message was not undoubtedly convincing until recent years when satellite-observed GRACE data became available. Using a geodesy approach, Han et al. (2009) employed a fixed-velocity version of TRIP in the Amazon river basin and its vicinity, and compared the model simulations to the residual of GRACE raw measurements derived from removing all the gravity-influencing factors except for the horizontally moving water. They demonstrated that the optimal flow velocity of TRIP in the Amazon varies between rising and falling water levels.

2.01.5.2 Global Changes and Global Hydrology

400 200 l G

lo

ba

e Ic

id hu

m

id ry

ic op

Ve

al

H

hu

um

m

id

id

Tr

(b)

um i-h

m Se

Se

m

ia

Ar

rid

id

0

Figure 10 Components of (a) the surface hydrologic balance, and (b) total evapotranspiration estimated under the second phase of the Global Soil Wetness project. From Dirmeyer PA, Gao XA, Zhao M, Guo ZC, Oki T, and Hanasaki N (2006) GSWP-2 multimodel anlysis and implications for our perception of the land surface. Bulletin of the American Meteorological Society 87: 1381–1397.

river discharge (Oki et al., 1999). The first version of TRIP adopted a primitive fixed-velocity scheme (Miller et al., 1994), while the variable-velocity version was also developed (NgoDuc et al., 2007). TRIP was coupled in some general circulation model (GCM) projections used in the Intergovernmental Panel on Climate Change (IPCC) Assessment Report 4 (AR4) to identify the impact of climate change on hydrological cycles (Falloon and Betts, 2006), and there have been some studies of future assessment on the world water resources and global flood disasters utilizing the TRIP model, as well (Oki and Kanae, 2006; Hirabayashi and Kanae, 2009). Further, Kim et al. (2009) underscored the importance of river component in terrestrial water storage (TWS) variation over global river basins. To reduce simulation uncertainty, ensemble simulations were performed with multiple precipitation data, and a localized Bayesian model averaging technique was applied to TRIP simulation. Figure 11 shows that river storage not only explains different portions of total TWS variations, but also plays different roles in different climatic regions. It is the most dominant water-storage component in wet basins (e.g., the Amazon) in terms of amplitude, and it acts as a buffer which smoothes the seasonal variation of total TWS especially in snow-dominated basins (e.g., the Amur). It signifies that model simulation of TWS may not be able to properly reproduce the amplitude and

Macroscale hydrological models have also been developed in response to the societal expectations for solving current and future world water issues. There has been a concrete demand for the information on how much water resources are available now and what kinds of changes are projected for the future. Conventionally, available freshwater resources are commonly defined as annual runoff estimated by historical river-discharge data or water-balance calculation (Lvovitch, 1973; Baumgartner and Reichel, 1975; Korzun, 1978). Such an approach has been used to provide valuable information on the annual freshwater resources in many countries. Atmospheric water balance using the water-vapor flux-convergence data could be alternatively used to estimate global distribution of runoff owing to the advent of atmospheric reanalysis and data-assimilation system (Oki et al., 1995). Simple analytical water-balance models have been widely used to estimate global-scale available freshwater resources in the world since the beginning of this century (Alcamo et al., 2000; Vo¨ro¨smarty et al., 2000; Do¨ll et al., 2003; Rockstro¨m et al., 2009). Later, LSMs were used to simulate global water cycles (Oki et al., 2001; Dirmeyer et al., 2006). Changes of hydrological cycles during the twenty-first century associated with anticipated climate change are projected (Milly et al., 2005; Nohara et al., 2006), and their impacts on the demands and supplies of global water resources are estimated assuming future climatic and social-change scenarios (Arnell, 2004; Alcamo et al., 2007; Shen et al., 2008). Some of those estimations were calibrated by multiplying an empirical factor for the river basins, where observed river-discharge data are available. However, recent model simulations with advanced climate forcing data can estimate global runoff distribution with adequate accuracy without the need of calibration (Hanasaki et al., 2008a). Changes in extreme river discharge are also of interest now (Hirabayashi et al., 2008). Several recently developed macroscale hydrological models for water-resource assessment also include a reservoir-operation scheme (Haddeland et al., 2006; Hanasaki et al., 2006) in order to simulate the real hydrological cycles, which are significantly influenced by anthropogenic activities and modified from natural hydrological cycles even on the global

20

Global Hydrology Discharge (mm per month)

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Figure 11 (a) Seasonal variations of gauged discharge (black solid line), discharge routed by TRIP (red solid line) and runoff without routing (gray dashed line). (b) Seasonal variations of GRACE observed TWSA (black solid line), simulated TWSA with river storage (red solid line), simulated TWSA without river storage (gray dashed line), and the major water storage components in TWS. Gray crosses ( þ ), green circles (K), and blue triangles (m) represent the individual storage component of snow water, soil moisture, and river storage, respectively. (c) Interannual variations of relative TWS: the GRACE observation (black dot), simulation with river storage (red solid line), and simulation without river storage (gray dashed line). Each area shaded by blue, gray, and green indicates the portion of river storage, snow water, and soil moisture in the simulated relative TWS, respectively. From Kim H, Yeh P, Oki T, and Kanae S (2009) The role of river storage in the seasonal variation of terrestrial water storage over global river basins. Geophysical Research Letters 36: L17402 (doi:10.1029/2009GL039006).

scale in the Anthropocene (Crutzen, 2002). An integrated water-resources model is further coupled with a crop-growth submodel, which can simulate the timing and quantity of irrigation requirement, and a submodel, which can estimate environmental flow requirement (Hanasaki et al., 2008a). Such an approach is able to assess the balances of water demand and supply on a daily timescale, and a gap in the subannual distribution of water availability and water use can be detected in the Sahel, the Asian monsoon region, and southern Africa, where conventional water-scarcity indices such as the ratio of annual water withdrawal to water availability and available annual water resources per capita (Falkenmark and Rockstro¨m, 2004) cannot properly detect the stringent balance between demand and supply (Hanasaki et al., 2008b).

2.01.5.3 Global Trace of Water Cycles Numerical models can be associated with a scheme tracing the origin and flow path as if tracing the isotopic ratio of water (Yoshimura et al., 2004; Fekete et al., 2006). Such a flow-tracing function of water in the integrated water-resources model (Hanasaki et al., 2008a) considering the sources of water withdrawal from stream flow, medium-size reservoirs, and nonrenewable groundwater, in addition to precipitation to croplands, enabled the assessment of the origin of water producing major crops (Hanasaki et al., 2010). Figure 12(a) illustrates the ratio of blue water to total evapotranspiration during cropping period in irrigated croplands. Here, the blue water is defined as that part of evapotranspiration originating from irrigation, whereas the green water is from precipitation (see Falkenmark and Rockstro¨m, 2004). Figure 12(a) shows a distinctive geographical distribution in the dependence on blue water. In addition, the ratios of the source of blue water for stream flow including the influence of large reservoirs,

medium-size reservoirs, and nonrenewable and nonlocal blue water are shown in Figures 12(b)–12(d). Areas highly dependent on nonrenewable and nonlocal blue water were detected in Pakistan, Bangladesh, western part of India, north and western parts of China, some regions in the Arabian Peninsula, and the western part of the United States through Mexico. Cumulative nonrenewable and nonlocal blue-water withdrawals estimated by the model correspond fairly well with the country statistics of total groundwater withdrawals (Hanasaki, 2009, personal communication), and such an integrated model has the ability to quantify the global virtual water flow (Allan, 1998; Oki and Kanae, 2004) or water footprint Hoekstra and Chapagain, 2007) through major crop consumption (Hanasaki et al., 2010). It is apparent that these achievements illustrate how the framework of global off-line simulation of LSMs, coupled with lateral river-flow model and/or anthropogenic activities, driven by the best-available meteorological forcing data, such as precipitation and downward radiation, is relevant for estimating global energy and water cycles, validating the estimates and sometimes the quality of forcing data with independent observations, and improving the models themselves. There are attempts to utilize this framework for assessing the impacts of climate change on future hydrological cycles which would demand adaptation measures in waterresources management, flood management, and food production. For such purposes, it is necessary to develop reliable forcing data for the future, based on GCM projections probably with bias corrections and spatial and temporal downscaling, as well as developing best estimates for the future boundary conditions for hydrological simulations such as vegetation type and land use/land cover. Figure 13 summarizes the concept of how the forcing data have a large impact on the accuracy of the output from theories, equations, and numerical models. Certainly, the spatial

Blue/(blue+green)

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Figure 12 (a) The ratio of blue water to the total evapotranspiration during a cropping period from irrigated cropland (the total of green and blue water). The ratios of (b) streamflow, (c) medium-size reservoirs, and (d) nonrenewable and nonlocal blue-water withdrawals to blue water. From Hanasaki N, Inuzuka T, Kanae S, and Oki T (2010) An estimation of global virtual water flow and sources of water withdrawal for major crops and livestock products using a global hydrological model. Journal of Hydrology 384: 232–244.

22

Global Hydrology

and temporal boundary conditions as well as the field information characterizing the target region, such as land cover and land use, are critically important to obtain reasonable estimates. Therefore, it is recommended to examine the quality of forcing data and boundary conditions along with revising the core theory, principle equations, and model code, or tuning model parameters, for hydrological modeling. It is particularly important for global-scale studies since uncertainties in the forcing data and field information are relatively large compared to those at the local scale. It should also be recalled that the applicability and accuracy of the model are highly dependent on the specific temporal and spatial scales.

Owing to recent advancements in global earth-observation technology and macroscale modeling capacity, global hydrology can now provide basic information on the regional hydrological cycle which may support the decision-making process in the integrated water-resources management. It should also be examined to what extent such a framework of off-line simulation of LSMs can be applied to finer spatial and temporal scales, such as 1-km grid spacing and hourly time interval. For such research efforts, observational data from regional studies can provide significant information, and efforts to integrate data sets from various regional studies should be promoted.

2.01.5.5 Research Opportunities in Global Hydrology 2.01.5.4 Interactions of Global- and Local-Scale Hydrology As described in Oki et al. (2006), water, as one of the major components of the global climate system, is one of the major cross-cutting axes in the Earth system science. The water cycle transports various materials, such as sediments and nutrients, from land to the oceans. Water resources are closely related to energy, industry, and agricultural production. Of course, water is indispensable for life and supports health. Water issues are related to poverty, and providing access to safe drinking water is one of the key necessities for sustainable development. In the past, water issues remained local issues; however, due to the increase in international trade and mutual interdependence among countries, water issues now often need to be dealt with on the global scale, and require information on global hydrology for their solutions. Sharing hydrological information relating to the transboundary river basins and shared aquifers will help reduce conflict between relevant countries, and quantitative estimates of recharge amounts or potentially available water resources will assist in implementing sustainable water use. Global hydrology is not merely concerned with global monitoring, modeling, and world water-resources assessment.

There are still challenging scientific issues to be resolved in global hydrology. For instance, separation of rain and snow on the global scale, as illustrated in Figure 1, is of interest. However, this is based on numerical-model estimates and it is quite uncertain about the accuracy of the numbers or even the ratio between rain and snow (here, approximately 8:1). The situation is similar for the ratio between surface and sub-surface runoff (here, approximately 1:2). It is reasonable to infer direct groundwater discharge from land to oceans to be a residual between total runoff estimated by numerical models subtracted by observation-based total discharge; however, the accuracy of such an estimate is yet to be determined. As described in Section 2.01.1, it is believed that total amount of the water on the Earth is conserved on a timescale shorter than geological timescales; however, do we have any reliable observational evidence for it? Our knowledge seems to be incomplete on the water exchange between Earth’s surface and mantle, although a recent report suggested that the lower mantle may store 5 times more water than the ocean (Murakami et al., 2002). The direct groundwater discharge to the ocean, estimated to be about 10% of total river discharge globally (Church, 1996),

Sources of uncertainties • Quantity • Runoff, ET, … • Quality • Solutes • Isotopic ratio • Sediments, …

Environmental info. Precip., radiation, temp., humidity, wind,… (atmospheric forcing)

Regional info. soil, vegetation, topography, basin area, … (parameters)

Theory, equation, or model

Initial/boundary conditions Applicability and accuracy would differ by temporal and spatial scales, and the characteristics of the target region.

Output with various temporal scales • Annual mean • Daily, hourly, … • Duration curve, … • Extremes, …

Figure 13 Sources and causes of uncertainties and errors in the hydrological simulations and estimations.

Global Hydrology

is included in the river discharge plotted in Figure 1. According to the model estimates by Koirala (2010), groundwater recharge in the grid boxes of 11 11 in longitude and latitude near the coastal line is totally 5890 km3 yr1, and it is approximately 13% of the annual total runoff from the continents to oceans of 45 500 km3 yr1. Even though the total amount of groundwater recharge depends on the grid size defining the coastal region, is it just a coincidence or is it that groundwater recharge in the coastal areas provides a good proxy of direct groundwater runoff from continents to oceans? Scientific interest in global hydrology has increased since the 1980s as public awareness of global environmental issues and process interactions, such as El Nin˜o events and anthropogenic climate change, has risen. Further work is required for the detection and attribution of the present-day hydrological changes – in particular, changes in the intra-seasonal water availability and in the frequency and magnitude of extreme events. Uncertainties in the future projections of hydrological quantities and water qualities should be reduced and quantitative accuracy should be enhanced, particularly for runoff regime changes. It is further anticipated that feedbacks of mitigation and adaptation measures for the concerned climate change on water sectors will be assessed for proper policymaking.

23

In addition to anthropogenic climate change, it is of interest in global hydrology to assess the impact of land-use change, such as deforestation and urbanization, human activities, such as reservoir construction and water withdrawals for irrigation, industry, and domestic water uses, and emission of air pollutants which would have been suppressing weak rainfall and modulate precipitation occurrence weekly.

2.01.5.6 Research Gaps in Global Hydrology Hydro-meteorological monitoring networks need to be maintained and further expanded to enable the analysis of hydro-climatic trends at the local level and the improvement in the accuracy of predictions, forecasts, and early warnings. As clearly illustrated in Figure 14 (from Oki et al., 1999), global hydrological simulations are relatively poor in areas with little in-situ observations. Basic observational networks on the ground are critically indispensable for proper monitoring and modeling of global hydrology; however, it is also required to utilize remotely sensed information in order to fill the gaps of in-situ observations. One of the current trends in the utilization of remote sensing technique is the so-called data

Runoff estimation error (mean 11 LSMs) and the density of raingauges in each river Basin Y1987 600

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Runof (model observation) (mm yr−1) Figure 14 Comparisons between the density of rain gauge (/106 km2) used in preparing the forcing precipitation and the mean bias error (mm yr1) of 11 LSMs for 150 major river basins in the world in 1987 and 1988. From Oki T, Nishimura T, and Dirmeyer P (1999) Assessment of annual runoff from land surface models using total runoff integrating pathways (TRIP). Journal of the Meteorological Society of Japan 77: 235–255.

24

Global Hydrology

assimilation, which optimally merges observations (not limited to remote sensing) with numerical-model estimates. Reliable observational data are essentially necessary not only as the forcing data for global hydrological modeling, but also for the validation of model estimates. River discharge and soil-moisture data are critically important for global hydrological studies. However, contributions from the operational agencies in the world are not yet well established and need to be enhanced. Isotopic ratio of rainwater is collected and the data set is available through the Global Network of Isotopes in Precipitation (GNIP) by the International Atomic Energy Agency (IAEA). These observational data can be used to investigate the routes and mechanisms of how the evaporated waters from the ocean surface are transported and precipitated at particular locations, which can be estimated by water-vapor transfer models with the consideration of isotopic processes (Yoshimura et al., 2008). The information on the isotopic ratio of river waters is not well organized and cannot be used easily on the global scale. Current global hydrological modeling has not yet integrated most of the latest achievements in process understanding and regional- or local-scale modeling studies. Global simulation of solutes and sediments are emerging. Both natural and anthropogenic sources should be considered, as for nutrients, and probably such models should be coupled with agricultural models which simulate crop growth. Moreover, it is rather difficult that river ice jams can be simulated properly by current hydrologic models. For both problems of water quality and ice jams, a proper simulation of water temperature in rivers and lakes are requisite. Moreover, the representation of groundwater has been rather simple in global hydrological modeling. Some of the above issues have not been emphasized in the current global hydrological modeling due to their relatively minor impact on the climatic feedbacks from the land surface to the atmosphere. It is meaningful to recall that global hydrology has been developed in cooperation with global climate modeling; however, it is time to develop global hydrological models primarily for responding to the demands of understanding the hydrological cycle on the land surface and for supporting better water-resources management. From this point of view, integrated hydrological and water-resource models, which consider natural and anthropogenic water cycles and are coupled with crop models and reservoir operation models in order to provide a more realistic impact assessment and support the design of practical adaptation measures, should be developed and implemented.

Acknowledgments This study was funded by Grants-in-Aid for Scientific Research from the Japan Society for the Promotion of Science (19106008 4), the Global Environment Research Fund of the Ministry of Environment (S-5), and the Innovation Program of Climate Change Projection for 21st Century of the Ministry of Education, Culture, Sports, Science, and Technology of Japan.

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2.02 Precipitation D Koutsoyiannis, National Technical University of Athens, Athens, Greece A Langousis, Massachusetts Institute of Technology (MIT), Cambridge, MA, USA & 2011 Elsevier B.V. All rights reserved.

2.02.1 2.02.1.1 2.02.1.2 2.02.1.3 2.02.1.4 2.02.1.5 2.02.1.5.1 2.02.1.5.2 2.02.1.5.3 2.02.1.5.4 2.02.1.5.5 2.02.2 2.02.2.1 2.02.2.2 2.02.2.3 2.02.2.3.1 2.02.2.3.2 2.02.2.4 2.02.2.4.1 2.02.2.4.2 2.02.2.5 2.02.2.5.1 2.02.2.5.2 2.02.2.5.3 2.02.2.5.4 2.02.2.5.5 2.02.2.5.6 2.02.2.5.7 2.02.3 2.02.3.1 2.02.3.1.1 2.02.3.1.2 2.02.3.1.3 2.02.3.2 2.02.3.2.1 2.02.3.2.2 2.02.3.3 2.02.3.3.1 2.02.3.3.2 2.02.3.3.3 2.02.4 2.02.4.1 2.02.4.2 2.02.4.3 2.02.4.4 2.02.4.5 2.02.5 2.02.5.1 2.02.5.2 2.02.5.3 References

Introduction The Entrancement of Precipitation Forms of Precipitation Precipitation Metrics The Enormous Variability of Precipitation Probability and Stochastic Processes as Tools for Understanding and Modeling Precipitation Basic concepts of probability Stochastic processes Stationarity Ergodicity Some characteristic stochastic properties of precipitation Physical and Meteorological Framework Basics of Moist Air Thermodynamics Formation and Growth of Precipitation Particles Properties of Precipitation Particles Terminal velocity Size distribution Clouds and Precipitation Types Cumulus cloud systems Stratus cloud systems Precipitation-Generating Weather Systems Fronts Mechanical lifting and orographic precipitation Extratropical cyclones Isolated extratropical convective storms Extratropical squall lines and rainbands Monsoons Tropical cyclones Precipitation Observation and Measurement Point Measurement of Precipitation Measuring devices Typical processing of rain gauge data Interpolation and integration of rainfall fields Radar Estimates of Precipitation Basics of radar observation and measurement Radar observation of distributed targets and the estimation of precipitation Spaceborne Estimates of Precipitation The IR signature of cloud tops The visible reflectivity of clouds The microwave signature of precipitation Precipitation modeling Rainfall Occurrence Rainfall Quantity Space–Time Models Rainfall Disaggregation and Downscaling Multifractal Models Precipitation and Engineering Design Probabilistic versus Deterministic Design Tools Extreme Rainfall Distribution Ombrian Relationships

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2.02.1 Introduction 2.02.1.1 The Entrancement of Precipitation Precipitation and its related phenomena, such as cloud formation and movement, thunders, and rainbow, are spectacular (Figure 1) due to their huge diversity and complexity. This complexity makes them difficult to comprehend, model, and predict. Hence, it is understandable that ancient civilizations explained these phenomena in a hyperphysical manner, assuming that deities were responsible for their creation. For example, in Greek mythology, some of the phenomena were deified (e.g., Iris is the name of a goddess as well as of the rainbow), whereas the most impressive among them, thunders in particular, were attributed to the action of the King of the Gods, Zeus (Jupiter in Roman mythology; similar deities are Indra in Hinduism, Thor in Norse, etc.). Demystification of these processes and formation of the physical concept of the hydrological cycle was closely related to the birth of science, by the turn of the seventh century BC. While the hydrological cycle was founded as a concept in the sixth century BC by

Anaximander, Anaximenes, and Xenophanes, and was later advanced by Aristotle (Koutsoyiannis et al., 2007), certain aspects related to precipitation can be understood only within the frame of modern science. The fact that a solid or liquid hydrometeor resists gravity and remains suspended in the atmosphere in a cloud is counterintuitive, and needs advanced knowledge of physics, fluid dynamics, and statistical thermodynamics to be understood and modeled. The complexity of the processes involved in precipitation and their enormous sensitivity to the initial conditions (where tiny initial differences produce great differences in the final phenomena), retain, to this day, some of the ancient mythical and magical magnificence of the societal perception of precipitation. People still believe in hyperphysical interventions in matters concerning precipitation. As put by Poincare´ (1908), father of the notion of chaos: Why do the rains, the tempests themselves seem to us to come by chance, so that many persons find it quite natural to pray for rain or shine, when they would think it ridiculous to pray for an eclipse?

Figure 1 Precipitation and related phenomena (from upper-left to lower-right): Monsoon rainfall (Pune, India, September 2009; photo by D Koutsoyiannis); snowy mountainous landscape (Mesounta, Greece, December 2008; from http://www.mesounta.gr/mesounta/ist_eik1/ 07_xion_03.htm); thunder (Athens, Greece, November 2005; from the photo gallery of Kostas Mafounis); rainbow (Mystras, Greece, April 2008, from laspistasteria.wordpress.com/2008/04/08/rainbow-3/).

Precipitation

Amazingly, however, and at the very same time, there is little disbelief in some climate modelers’ prophecies (or outputs of global circulation models (GCMs)) of the precipitation regimes over the globe in the next 100 years or more. This indicates an interesting conflict between perceptions of precipitation – that it is so unstable, uncertain, and unpredictable that prayers are needed to invoke precipitation, and that for some scientists, the future evolution of precipitation on Earth is still predictable in the long term. The latter belief concerns not only the general public, but also the scientific community. For example, a Google Scholar search with either of the keywords ‘precipitation’ or ‘rainfall’, plus the keywords ‘climate change’ and ‘GCM’, locates 21 700 publications (as of August 2009), of which about 200 have been cited 100 times or more. This huge list of results appears despite the fact that climate modelers themselves admit to the performance of their models being low, as far as precipitation is concerned (Randall et al., 2007). An independent study by Koutsoyiannis et al. (2008), which compares model results for the twentieth century with historical time series, has shown that the models are not credible at local scales and do not provide any basis for assessment of future conditions. These findings demonstrate that, even today, the perception of precipitation, not only by the general public, but even by scientists specialized in the study of precipitation, meteorologists, climatologists, and hydrologists, continues to be contradictory, problematic, and, in some sense, mysterious.

2.02.1.2 Forms of Precipitation Precipitation occurs in a number of forms, either liquid or solid, or even mixed (sleet). Liquid precipitation includes rainfall and drizzle, where the former is the most common and most significant, and the latter is characterized by much smaller drop sizes and lighter intensity. Dew is another liquid form, formed by condensation of water vapor (mostly at night) on cold surfaces (e.g., on tree leaves). Most important among the solid forms of precipitation are snow and hail. At high latitudes or at high altitudes, snow is the predominant form of precipitation. Snowfall may occur when the temperature is low and snow accumulates on the ground until the temperature rises sufficiently for it to melt. On the other hand, hail may fall in relatively high temperature and usually melts rapidly. While hailstones are amorphous and usually large (one to several centimeters in diameter), snowflakes are symmetrical and visually appealing with a tremendous variety of shapes, so that no two snowflakes are the same. Occult precipitation is induced when clouds or fog is formed in forested areas, and it includes liquid (fog drip) and solid (rime) forms. Fog drip occurs when water droplets are deposited on vegetative surfaces, and the water drips to the ground. Rime is formed when supercooled air masses encounter exposed objects, such as trees, that provide nucleation sites (see Sections 2.02.2.1 and 2.02.2.2) for formation and buildup of ice, much of which may fall to the ground in solid or liquid form. In some places (e.g., in humid forested areas), precipitation of this type may reach significant amounts; for example, rime constitutes about 30% of the annual precipitation in a Douglas fir forest in Oregon (Harr, 1982; Dingman,

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1994) and about 30% of total precipitation in fir-forested mountainous areas of Greece (Baloutsos et al., 2005), and it is the sole precipitation type on the rainless coast of Peru (Lull, 1964; Dingman, 1994).

2.02.1.3 Precipitation Metrics The principal metric of precipitation is the rainfall depth h (commonly expressed in millimeters) that falls at a specified point in a specified period of time t; this can be easily perceived and measured by a bucket exposed to precipitation. A derivative quantity is the precipitation intensity

i :¼

dh dt

ð1Þ

with units of length divided by time (typically mm h1, mm d1, and mm yr1). Since it cannot be measured directly (at an instantaneous time basis), it is typically approximated as

i¼

Dh Dt

ð2Þ

where Dh is the change of the depth in a finite time interval Dt. The intensity derived from Equation (2) is a time-averaged value – but at a point basis. Spatial averaging at various scales is always very useful as can be seen in Section 2.02.1.4. This averaging needs precipitation measurement at several points, followed by appropriate numerical integration methods (see Section 2.02.3.1). While traditional precipitation-measurement networks are sparse, thus making the estimation of areal precipitation uncertain, in recent decades, new measurement techniques have been developed implementing radar and satellite technologies (Sections 2.02.3.2 and 2.02.3.3). These provide a detailed description of the spatial distribution of precipitation, thus enabling a more accurate estimation. The latter techniques inherently involve the study of other metrics of precipitation such as the distribution of the size, velocity, and kinetic energy of the precipitation particles, and the socalled radar reflectivity (Section 2.02.3.2). Furthermore, the quantitative description of the processes related to the fall, accumulation, and melting of snow involves a number of additional metrics, such as the snowfall depth (new snow falling), the snowcover depth or snowpack depth (the depth of snow accumulated at a certain point at a particular time), the snow density rs, and the water equivalent of snowfall or of snowpack, defined as h ¼ h0 rs/rw, where h0 is the snowfall or the snowcover depth, and rw ¼ 1000 kg m3 is the liquid water density. Typical values of rs for snowfall range between 0.07 and 0.15rw (e.g., Dingman, 1994) but a commonly used value is rs ¼ 0.1rw ¼ 100 kg m3. For this value, a snowfall depth of say, 10 cm, corresponds to a precipitation water equivalent of 10 mm. The density of snowpack is generally larger than 0.1rw (because of compaction due to gravity and other mechanisms) and depends on the elapsed time and snowpack depth. After a few days, it is about 0.2rw, whereas after some months it may become about 0.4rw.

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2.02.1.4 The Enormous Variability of Precipitation The different phases and forms of precipitation, and the different shapes of precipitation particles (drops, flakes, and hailstones) are just a first indicator of the great diversity of the precipitation phenomena. At a macroscopic level and in quantitative terms, this diversity is expressed by the enormous variability of the precipitation process, in space and time, at all spatial and temporal scales. Intermittency is one of the aspects of variability, but even in areas or time periods in which precipitation is nonzero, the precipitation depth or average intensity is highly variable. Figure 2 shows the spatial variability of precipitation over the globe in mm d1 at the climatic scale (average for the 30year period 1979–2008) and at an annual scale (average for the year 2006), based mostly on satellite data (see Figure 2 caption and Section 2.02.3.3). While the average precipitation rate over the globe and over the specified 30-year period is 2.67 mm d1 or 977 mm yr1, we observe huge differences in different areas of the globe. In some areas, mostly in tropical seas and in equatorial areas of South America and Indonesia, this rate exceeds 10 mm d1 or 3.65 m yr1. On the other hand, in large areas in the subtropics, where climate is dominated by semi-permanent anticyclones, precipitation is lower than 1 mm d1 or 365 mm yr1. Significant portions of these areas in Africa, Australia, and America are deserts, where the average precipitation is much lower than 1 mm d1. In addition, in polar regions, where the available atmospheric moisture content is very low due to low temperature (see Section 2.02.2.1 and Figure 14), the amounts of precipitation are very small or even zero. For example, it is believed that certain dry valleys in the interior of Antarctica have not received any precipitation during the last 2 million years (Uijlenhoet, 2008). Figure 3 depicts the zonal precipitation profile and shows that the climatic precipitation rate at an annual basis is highest at a latitude of 51 N, reaching almost 2000 mm yr1 and has a second peak of about 1500 mm yr1 at 51 S. Around the Tropics of Cancer and Capricorn, at 23.41 N and S, respectively, the rainfall rate displays troughs of about 600 mm yr1, whereas at mid-latitudes, between 351 and 601 both N and S, rainfall increases again and remains fairly constant, close to the global average of 977 mm yr1. Then, toward the poles, it decreases to about 100 mm in Antarctica and slightly more, to 150 mm yr1, in the Arctic. Figure 3 also shows monthly climatic profiles for the months of January and July. It can be seen that the rainfall conditions for the 2 months are quite different, with the largest differences appearing at about 151 N and S and the smallest at about 301 N and S. Below 301 in the Northern Hemisphere, as well as above the Arctic Circle (66.61), rainfall is higher during summer (July) than during winter (January), but at mid-latitudes, this relationship is reversed. Similar conditions are met in the Southern Hemisphere (where January and July are summer and winter months, respectively). In both Figures 2 and 3, apart from climatic averages, the specific values for a certain year, namely 2006, are also shown. We observe that there are differences in the climatic values, manifesting temporal variability over the different years. This variability seems to be lower in comparison to the spatial

variability over the globe, as well as to the seasonal variability reflected in the profiles of different months. However, while the spatial variability over the globe and the seasonal variability are well comprehended and roughly explainable in terms of basic physical and astronomical knowledge (i.e., solar radiation, relationship of temperature and atmospheric moisture content, and motion of Earth), in other words, they are regular, the interannual variability is irregular, and difficult or even impossible to predict. Such irregular variability appears at finer timescales as well as at finer spatial scales. In fact, as easily understood from elementary statistics, as the spatial and/or temporal scale becomes finer, the variability increases. Figure 4 demonstrates how the variability of the spatial distribution of rainfall at a monthly temporal scale (January 2006) increases when the spatial scale decreases from 2.51  2.51 (upper panel) to 0.251  0.251 (lower panel). Clearly, the areas of equal rainfall amount (including areas of negligible rainfall, i.e., o1 mm d1E0.04 mm h1), which are smooth in the upper panel become rough and erratic in the lower panel. Moreover, the maximum observed rainfall is 21 mm d1 (monthly 651 mm) in the upper panel and 1.2 mm h1 (monthly 893 mm) in the lower panel. Figure 5 demonstrates the increasing variability with the decreasing timescale. Specifically, it depicts how the image of the rainfall distribution changes at a daily scale (9 January 2006) and at a sub-daily scale, at 3-hourly intervals of the same day. The differences between Figure 5 and Figure 4 are prominent. Especially at the 3-hourly scale, a vast part of the globe receives no rainfall, and the part that receives rainfall, it is irregularly distributed, yet not showing a totally random pattern. The maximum observed rate during this 3-hourly interval is 22 mm h1, about 18 times higher than the maximum rate at the monthly scale shown in Figure 4. The lowest panels of Figure 5 provide a zoom-in over the area lying between 91 N–51 S and 78–921 E, which is located in the Indian Ocean, south-east of Sri Lanka. This area received a large amount of rainfall on this particular day, with a rate that is nonuniform in space and time. Figures 6–8 focus on the temporal variability of precipitation. Figure 6 depicts the monthly and annual variation of the average precipitation over the globe. We can see that at both scales, the variability is remarkable. Thus, the annual precipitation in the last 30 years has varied between 957 and 996 mm and obviously much higher variation should have occurred in the past – but data of this type covering the entire globe do not exist for earlier periods. However, we can get an idea of earlier variation using rain gauge data (see Section 2.02.3.1) at certain locations. Perhaps, the oldest systematic observations of rainfall quantity in the world were made in Korea, in the fifteenth century. Rainfall records for the city of Seoul (37.571 N, 126.971 E, and 85 m) exist for the period since 1770, and are considered to be reliable (Arakawa, 1956; Wang et al., 2006, 2007). The recorded annual rainfall in Seoul is plotted in Figure 7 along with the running climatic averages at 10-year and 30-year timescales. The data are now available at a monthly scale from the climatic database of the Dutch Royal Netherlands Meteorological Institute (KNMI), while the monthly data for 1770–1907 appear also in Arakawa (1956).

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Figure 2 Precipitation distribution over the globe in mm d , (upper) at a climatic scale (average for the 30-year period 1979–2008) and (lower) at an annual scale (average for year 2006). Data and image generation due to the Global Precipitation Climatology Project (GPCP) made available by NASA at http://disc2.nascom.nasa.gov/Giovanni/tovas/rain.GPCP.2.shtml; resolution 2.51  2.51.

Comparisons show that the two time series are generally consistent, but not identical. The more modern data series has a few missing values, which generally correspond to high values of the older version (and it has been common practice in hydrometeorological data processing to delete very high

values or outliers, which are regarded suspect, see Section 2.02.3.1.2). In the time series plotted in Figure 7, these gaps have been filled in using the values of the older time series, and a few other missing values have been filled in with the average of the four nearest monthly values of the same month

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Precipitation (mm d−1) Figure 3 Zonal precipitation profile: precipitation averaged over all longitudes for different latitudes varying from 901N to 901S (–901N); data from GPCP.

(see justification in Section 2.02.3.1.3). The plot shows that during the 238 years of record, the annual rainfall varied between 634 and 3057 mm and the climatic 30-year average varied between 1139 and 1775 mm. These figures indicate a huge variability: the maximum observed annual rainfall is almost 5 times greater than the minimum and the maximum 30-year climatic rainfall is 55% higher than the minimum. Such observed changes underscore the ever-changing character of climate, and render future changes of precipitation predicted by climate modelers (which typically vary within 10– 20%; compare Fig. 10.12, upper left panel, in Meehl et al., 2007, with Figure 2 herein) to be unrealistically low and too unsafe to support planning. Figure 7 also includes a plot of another long time series, for Charleston City, USA (32.791 N, 79.941 W, and 3 m); the record begins in 1835. This time series is also available at the KNMI database, and a few missing monthly values have been filled in by the average of the four nearest monthly values of the same month. Here, the annual rainfall varied between 602 and 1992 mm (3.3 times higher than minimum) and the climatic 30-year average varied between 1135 and 1425 mm (25% higher than minimum). Finally, Figure 8 depicts the time series of a storm measured at unusually high temporal resolution, that is, 10 s. This storm, with duration 96 790 s or about 27 h starting at 199002-12T17:03:39, is one of several storms that were measured at the University of Iowa using devices that support high

sampling rates (Georgakakos et al., 1994). Figure 8 also includes plots at 5-min and hourly timescales. The minimum intensity was virtually zero at all three scales, whereas the maximum rainfall intensity was 118.7, 38.9, and 18.1 mm h1 at timescales of 10 s, 5 min, and 1 h, respectively. As the mean intensity during the storm is 3.89 mm h1, these maximum values are 30, 10, and 4.6 times higher than the mean. This example highlights the spectacular variability of rainfall, particularly at fine timescales (see also Uijlenhoet and SempereTorres, 2006). As the total rainfall amount of this storm event only slightly exceeds 100 mm, it could be thought of as a rather modest event. Storms with amounts much higher than this are often recorded even in semi-dry climates and, obviously, the variability of rainfall intensity during such storms is even higher.

2.02.1.5 Probability and Stochastic Processes as Tools for Understanding and Modeling Precipitation The high variability and the rough and irregular patterns in observed fields and time series are much more prominent in precipitation than in other meteorological variables such as atmospheric pressure or temperature. High variability implies high uncertainty and, unavoidably, this affects predictability in deterministic terms. Considering weather prediction as an example, it is well known that the forecasts of atmospheric pressure and temperature are much more reliable than those

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Figure 4 Monthly rainfall distribution over the globe in January 2006 in mm h : (upper) data with resolution 2.51  2.51 from GPCP; (lower) data with resolution 0.251  0.251 from the Tropical Rainfall Measuring Mission (TRMM) and Other Rainfall Estimate (3B42 V6) archive, made available by NASA at http://disc2.nascom.nasa.gov/Giovanni/tovas/TRMM_V6.3B42.shtml.

of precipitation. Numerical weather prediction (NWP) uses current weather conditions as an input to mathematical models of the atmosphere, which solve the flow (Navier– Stokes) equations, the thermodynamic energy equation, the state equation of gases, and the equation for conservation of water vapor, over a grid covering the entire atmosphere. The processes related to cloud formation and precipitation (see Section 2.02.2.2) are less accurately represented in these models. While the continuous improvement of NWP models resulted in a considerable reduction of forecast errors on pressure and temperature, the improvement in the so-called quantitative precipitation forecast (QPF) has been slower (Olson et al., 1995). Further, although the advances in computing infrastructure permitted the increase in model resolution that leads generally to an improvement of precipitation forecasts, recently, many authors have highlighted the limitations of such an approach (e.g., Mass et al., 2002; Lagouvardos et al., 2003; Kotroni and Lagouvardos, 2004). The major advancement in QPF in the last decades was the abandonment of the pure deterministic approach, which seeks a unique prediction, and the adoption of a more probabilistic approach to precipitation forecast, based on earlier ideas of Epstein (1969) and Leith (1974). In this approach, known as ensemble forecasting, the same model produces many

forecasts. To produce different forecasts, perturbations are introduced, for example, in the initial conditions, and, because of the nonlinear dynamics with sensitive dependence on the initial conditions (e.g., Lorenz, 1963), these perturbations are magnified in time, thus giving very different precipitation amounts in a lead time of 1 or more days. The different model outputs can then be treated in a probabilistic manner, thus assigning probabilities to rainfall occurrence as well as to the exceedance of a specified rainfall threshold. In this manner, although the model uses deterministic dynamics, the entire framework is of the Monte Carlo or stochastic type. This method is satisfactory for a time horizon of forecast of a few days. In hydrology, this time horizon is relevant in realtime flood forecasting. However, in hydrological design, horizons as long as 50 or 100 years (the lifetimes of engineering constructions) are typically used. For such long horizons, the use of deterministic dynamics and of the related laborious models would not be of any help. However, a probabilistic approach is still meaningful – in fact the only effective approach – and, in this case, it can be formulated irrespective of the dynamics. Rather, the probabilistic approach should be based, in this case, on historical records of precipitation, such as those displayed in Figures 7 and 8. The reasoning behind neglecting the deterministic dynamics is

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Figure 5 Spatial rainfall distribution at daily and sub-daily scale: (upper) daily rainfall over the zone between 501N and 501S on 9 January 2006; (middle) 3-hourly rainfall at 09:00 on the same day; (lower left) zoom-in of the upper panel for daily rainfall in the Indian Ocean south-east of Sri Lanka (shown in figure); (lower right) zoom-in of the middle panel for 3-hourly rainfall for the same area. Data in mm h1 with resolution 0.251  0.251 from the TRMM 3B42 V6 archive.

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that, beyond a certain time horizon (which in precipitation is of the order of several days), even the simplest nonlinear systems tend to a statistical equilibrium state. In this state, the probability distribution of the system properties, conditioned on the initial state, is practically equal to the marginal (i.e., unconditional) probability distribution of the same properties (Koutsoyiannis, 2009). This equilibrium, which is different from the typical thermodynamic equilibrium, corresponds to the maximization of the entropy of the vector of random variables defining the system state.

2.02.1.5.1 Basic concepts of probability Probability is thus not only a mathematical tool to model precipitation uncertainty, but also a concept for understanding the behavior of precipitation. Probabilistic thinking provides insights into phenomena and their mathematical descriptions, which may not be achievable in deterministic terms. It should be recalled that, according to the Kolmogorov (1933) system, probability is a normalized measure, that is, a function P that maps sets (areas where unknown quantities lie) to real numbers (in the interval [0, 1]). Furthermore, a random variable x is a single-valued function of the set of all elementary events (so that to each event, it maps a real number) and is associated with a probability distribution function. The latter is defined as

Fx ðxÞ :¼ Pfx r xg

ð3Þ

where x is any real number, which should be distinguished from the random variable x. (Distinction of random variables from their values is usually done by denoting them with upper case and lower case letters, respectively. This convention has several problems – e.g., the Latin x and the Greek w, if put in

upper case, are the same symbol X – other texts do not distinguish the two at all, thus creating another type of ambiguity. Here, we follow a different convention, in which random variables are underscored and their values are not.) Fx ðxÞ is a nondecreasing function of x with the obvious properties Fx ðNÞ ¼ 0 and Fx ðþNÞ ¼ 1. For continuous random variables (as is, for instance, the representation of a nonzero rainfall depth), the probability that a random variable x would take any particular value x is Pfx ¼ xg ¼ 0. Thus, the question of whether one particular value (say x1 ¼10 mm, assuming that x denotes daily rainfall at a location) is more probable than another value (say x2 ¼10 m, which intuitively seems extremely improbable) cannot be answered in terms of the probability function P, as all particular values have probability equal to zero. The derivative of F, that is,

f x ðxÞ :¼

dFx ðxÞ dx

ð4Þ

termed the probability density function, can provide this answer, as the quantity f x ðxÞ dx is the probability that rainfall will lie in an interval of length dx around x. Apparently, then the ratio f x ðx1 Þ=f x ðx2 Þ equals the ratio of the probabilities at points x1 and x2. These rather simple notions allow quantification of uncertainty and enable the production of different type of predictions, which offer a concrete foundation of rational decisions for the design and management of water-resourses projects. This quantification is sometimes (mostly in Bayesian statistics) referred to as ‘probabilization of uncertainty’ that is meant to be the axiomatic reduction from the notion of unknown to the notion of a random variable (Robert, 2007).

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Figure 7 Annual precipitation time series in two of the stations with the longest records worldwide: (upper) Seoul, Korea; (lower) Charleston City, USA. Data from the database of the Dutch Royal Netherlands Meteorological Institute (KNMI; http://climexp.knmi.nl) and additional information as shown in text.

2.02.1.5.2 Stochastic processes In the study of rainfall variation in time, the notion and the theory of stochastic processes provide the necessary theoretical framework. A stochastic process is defined as an arbitrarily (usually infinitely) large family of random variables xðtÞ (Papoulis, 1991). In most hydrological applications, time is discretized using an appropriate time step d; for integer i, the average of the continuous time process xðtÞ from t ¼ (i  1)d to t ¼ i d, is usually denoted xi and forms a discrete time stochastic process. The index set of the stochastic process (i.e., the set from which the index t or i takes its values) can also be a vector space, rather than the real line or the set of integers. This is the case, for instance, when we assign a random variable (e.g., rainfall depth) to each geographical location (a two-dimensional (2D) vector space) or to each location and time instance (a 3D vector space). Stochastic processes with a multidimensional index set are also known as random fields.

A realization x(t) (or xi) of a stochastic process xðtÞ (or xi ), which is a regular (numerical) function of the time t (or a numerical sequence in time i), is known as a sample function. Typically, a realization is observed at countable time instances (and not in continuous time, even if the process is of continuous-time type). This sequence of observations is also referred to as a time series. Clearly then, a time series is a sequence of numbers, whereas a stochastic process is a family of random variables. (Unfortunately, a large body of literature does not make this distinction and confuses stochastic processes with time series.) The distribution and the density functions of the random variable xi, that is,

Fi ðxÞ :¼ Pfxi r xg;

f i ðxÞ :¼

dFi ðxÞ dx

ð5Þ

Precipitation 3

6

9

12

18

181

21

24

37

27 h

120 10-s scale 5-min scale

100 Rainfall intensity (mm h−1)

Hourly scale 80

60

40

20

0 0

10 000

20 000

30 000

40 000

50 000

60 000

70 000

80 000

90 000

Time (s) Figure 8 Time series of a storm in Iowa, USA, measured at the University of Iowa with temporal resolution of 10 s; time zero corresponds to 199002-12T17:03:39. From Georgakakos KP, Carsteanu AA, Sturdevant PL, and Cramer JA (1994) Observation and analysis of Midwestern rain rates. Journal of Applied Meteorology 33: 1433–1444.

are called, respectively ‘first-order distribution function’ and ‘first-order density function’ of the process. Likewise, the second-order distribution function is Fi1 i2 ðx1 ; x2 Þ ¼ Pfxi1 r x1 ; xi2 r x2 g and this can be generalized to define the nth-order distribution function. It should be recalled that the expected value of a function g of one, two, or more random variables is the integral of g multiplied by the density f, that is,

E½gðxi Þ :¼

ZN

gðxÞf i ðxÞ dx

N

E½gðxi1 ; xi2 Þ ¼

ZN ZN

gðx1 ; x2 Þf i1 i2 ðx1 ; x2 Þ dx1 dx2

2.02.1.5.3 Stationarity As implied by the above notation, in the general setting, the statistics of a stochastic process, such as the mean and autocovariance, depend on time i and thus vary with time. However, the case where these statistical properties remain constant in time is most interesting. A process with this property is termed a ‘stationary’ process. More precisely, a process is called ‘strict-sense’ stationary, if all its statistical properties are invariant with a shift in the time origin. That is, the distribution function of any order of xiþj is identical to that of xi . A process is called ‘wide-sense stationary’ if its mean is constant and its autocovariance depends only on time differences (lags), that is,

E½Xi  ¼ m;

N N

ð6Þ The use of square brackets in E[ ] and the random variables xi rather than their values x signifies the fact that the expected value is not a function of the real number x; rather, it depends solely on the distribution function associated with the random variable xi. Of particular interest are the cases where gðxi Þ ¼ xi, where E½xi  ¼: mi is the mean value of xi , and gðxi1 ; xi2 Þ ¼ ðxi1  mi1 Þðxi2  mi2 Þ, where E½ðxi1  mi1 Þðxi2  mi2 Þ ¼: Ci1 i2 is the process autocovariance, that is, the covariance of the random variables xi1 and xi2 . The process variance (the variance of the variable xi ), is a special case of the latter, that is, Var½xi  ¼ Cii , whereas the standard devipffiffiffiffiffiffi ation is the square root of the latter, that is, si :¼ Cii . Consequently, the process autocorrelation (the correlation coefficient of the random variables xi1 and xi2 ) is ri1i2: ¼ Ci1i2/ (si1 si2).

E½ðXiþj  mÞðXi  mÞ ¼ Cj

ð7Þ

Evidently, the standard deviation is constant too, that is, si ¼ s, and the autocorrelation is a function of the time lag only, that is, riþj, i ¼ rj. A strict-sense stationary process is also wide-sense stationary, but the reverse is not true. A process that is not stationary is called nonstationary. In a nonstationary process, one or more statistical properties depend on time. A typical case of a nonstationary process is the cumulative rainfall depth whose mean obviously increases with time. For instance, let us assume that the instantaneous rainfall intensity iðtÞ at a geographical location and period of the year is a stationary process, with a mean m. Let us further denote by hðtÞ, the rainfall depth collected in a large container (a cumulative rain gauge) at time t, and assume that at the time origin, t ¼ 0, the container is empty. Clearly E½hðtÞ ¼ mt. Thus hðtÞ is a nonstationary process. It should be stressed that stationarity and nonstationarity are properties of a stochastic process, not of a sample function

38

Precipitation

or time series. There is some confusion in the literature about this, as there are several studies that refer to a time series as stationary or nonstationary. As a general rule, to characterize a process as nonstationary, it suffices to show that some statistical property is a deterministic function of time (as in the above example of the cumulative rainfall), but this cannot be directly inferred merely from a time series. To understand this, let us consider the time series of annual rainfall in Seoul, plotted in the upper panel of Figure 7. Misled by the changing regime of precipitation at the climatic scale, as manifest in the plot of the 30-year average, it would be tempting to note (1) an increasing trend in the period 1770–90; (2) a constant climate with high precipitation during 1790–1870; (3) a decreasing trend between 1870 and 1900; and (4) a constant climate with low precipitation thereafter. It is then a matter of applying a fitting algorithm to determine, say, a broken-line type of function to the time series, which would be called a deterministic function of time. The conclusion would then be that the time series is nonstationary. However, this is a wrong ex-post argument, which interprets the long-term variability of the processes as a deterministic function. Had the function been indeed deterministic, it would also apply to future times, which obviously is not the case. Comparison with the previous example (cumulative rainfall), where the deterministic function E½xðtÞ ¼ mt was obtained by theoretical reasoning (deduction) rather than by inspection of the data, demonstrates the real basis of nonstationarity. Koutsoyiannis (2006b) has provided a more detailed study of this issue. Stochastic processes describing periodic phenomena, such as those affected by the annual cycle of the Earth, are clearly nonstationary. For instance, the daily rainfall at a mid-latitude location cannot be regarded as a stationary process. Rather, a special type of a nonstationary process, whose properties depend on time in a periodical manner (are periodic functions of time), should be used. Such processes are called ‘cyclostationary’ processes.

implies that the random variable has zero variance. This is precisely the condition that makes a process ergodic, a condition that does not hold true for every stochastic process.

2.02.1.5.4 Ergodicity

However, this law hardly holds in geophysical time series, including rainfall time series, whatever the scale is. This can be verified based on the examples presented in Section 2.02.1.4. A more plausible law is expressed by the elementary scaling (power-law) property

The concept of ergodicity (from the Greek words ergon, work; and odos, path) is central to the problem of determining the distribution function of a process from a single sample function (time series). A stationary stochastic process is ergodic if any statistical property can be determined from a sample function. Given that, in practice, the statistical properties are determined as time averages of time series, the above statement can be formulated alternatively – a stationary stochastic process is ergodic if time averages equal ensemble averages (i.e., expected values). For example, a stationary stochastic process is mean ergodic if

E½xi  :¼ limN-N

N 1X xi N i¼1

ð8Þ

The left-hand side in the above equation represents the ensemble average, whereas the right-hand side represents the time average, for the limiting case of infinite time. While the left-hand side is a parameter, rather than a random variable, the right-hand side is a random variable (as a sum of random variables). Equating a parameter with a random variable

2.02.1.5.5 Some characteristic stochastic properties of precipitation It has been widely accepted that rainfall exhibits some autocorrelation (or time dependence) if the timescale of study is daily or sub-daily, but this dependence vanishes at larger timescales, such as monthly or yearly. Thus, for timescales monthly and above, rainfall data series have been traditionally treated as independent samples. Mathematically, such a perception corresponds to a Markovian dependence at fine timescales, in which the autocorrelation decreases rapidly with time lag in an exponential manner, that is,

rj ¼ rj

ð9Þ

where r: ¼ r1. Then for a large lag j, or for a large scale of aggregation and even for the smallest lag (one), the autocorrelation is virtually zero (e.g., Koutsoyiannis, 2002). If xi denotes the stochastic process at an initial timescale, which is designated as scale 1, then the averaged process at an aggregated timescale k ¼ 2, 3, y, is ðkÞ

xi :¼

xði1Þkþ1 þ y þ xi k

ð10Þ

ð1Þ

(with xi  xi ). Let s(k) be the standard deviation at scale k. In processes xi independent of time, s(k) decreases with scale according to the well-known classical statistical law of inverse square-root, that is,

s s ðkÞ ¼ pffiffiffi k

s ðkÞ ¼

s k 1H

ð11Þ

ð12Þ

where H is the so-called Hurst exponent, named after Hurst (1951) who first studied this type of behavior in geophysical time series. Earlier, Kolmogorov (1940), when studying turbulence, had proposed a mathematical model to describe this behavior. This behavior has been known by several names, including the Hurst phenomenon, long-term persistence, and long-range dependence, and a simple stochastic model that reproduces it is known as a simple scaling stochastic model or fractional Gaussian noise (due to Mandelbrot and van Ness, 1968). Here, the behavior is referred to as the Hurst–Kolmogorov (HK) behavior or HK (stochastic) dynamics and the model, as the HK model. This behavior implies that the autocorrelation decreases slowly, that is, according to a power-type function,

Precipitation

with lag j: ðkÞ

ri ¼ rj ¼ ð1=2Þ½ðjj þ 1jÞ2H þ ðjj  1jÞ2H   jjj2H E Hð2H  1Þj2H2

ð13Þ

so that independence virtually never holds, unless H ¼ 0.5, a value which reinstates classical statistics including the law in Equation (11). Most often, natural processes including rainfall are positively correlated and H varies in the range (0.5, 1). The above framework is rather simple and allows easy exploration of data to detect whether they indicate consistence with classical statistics or with the HK behavior. A simple exploration tool is a double logarithmic plot of the estimates of standard deviation s(k) versus scale k, which is known as a ‘climacogram’. (ClimacogramoGreek Klımako´grammao(climax ´ (klımax) ¼ scale) þ (gramma (gramma) ¼ written).) In such a plot, the classical law and the HK law are manifested by a linear arrangement of points with slopes –0.5 and H  1, respectively. We must bear in mind, however, that a consequence of the HK law in Equation (12) is that the classical estimator of the variance

s2 ¼

n 1 X ðxi  xÞ2 n  1 i¼1

ð14Þ

ðnÞ

where n is the sample size, x  x1 is the estimator of the mean, and s the estimator of standard deviation, implies negative bias if there is temporal dependence. The bias becomes very high for HK processes with H approaching 1. Apparently then, s could be a highly biased estimator of s; an approximately unbiased estimator is (Koutsoyiannis, 2003a; Koutsoyiannis and Montanari, 2007):

rffiffiffiffiffiffiffiffiffiffiffiffiffi n0 s :¼ s n0  1

E

ð15Þ

39

where n0 is the equivalent (or effective) sample size, that is, the sample size that in the framework of classical statistics would ðnÞ lead to the same uncertainty (in the estimation of m by x  x1 ) as an HK series yields with sample size n. For an HK process, n0 is related to n by

n0 ¼ n 2ð1HÞ

ð16Þ

It can be seen that n0 can be very small even for high n if H pffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi ffi is high, and thus the correcting factor n0 =ðn0  1Þ in Equation (15) can be very large (see Koutsoyiannis and Montanari, 2007). Returning to the time series of globally averaged monthly precipitation in the 30-year period 1979–2008, which has been discussed earlier and is displayed in Figure 6, we may now study its statistical properties for several timescales. As the precipitation amounts are averaged over the entire globe, the effect of seasonality is diminished and the time series can be modeled using a stationary process rather than a cyclostationary one. Figure 9 depicts the climacogram, that is, a logarithmic plot of standard deviation versus scale. Empirical estimates of standard deviations have been calculated using both the classical estimator in Equation (14) and the HK estimator in Equation (15). Theoretical curves resulting from the classical statistical model (assuming independence), the Markovian model, and the HK model have also been plotted. For the Markovian model, the lag one autocorrelation coefficient, estimated from the monthly data, is r ¼ 0.256 and for the HK model, the estimate of the Hurst coefficient is H ¼ 0.70. This can be obtained readily from the slope of the straight line fitted to the group of empirical points in Figure 9, which should be H  1. Here, a slightly modified algorithm from Koutsoyiannis (2003a) has been used for the estimation of H. Overall, Figure 9 clearly demonstrates that the empirical

−1.2 Empirical, classical estimate Empirical, HK estimate Classical statistical model Markov model HK model

Log(standard deviation in mm d−1)

−1.3

−1.4

−1.5

−1.6

−1.7

−1.8 0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

Log(scale in months) Figure 9 Climacogram of the time series of globally averaged monthly precipitation in the 30-year period 1979–2008 shown in Figure 6. The estimate of the Hurst coefficient for the HK model is H ¼ 0.70.

40

Precipitation

points are inconsistent with the classical and Markovian models and justify an assumption of HK behavior. Similar plots have been constructed, and are shown in Figure 10, for the annual precipitation time series from Seoul, Korea, and Charleston City, USA, displayed in Figure 7. Again, the empirical evidence from data precludes the applicability of the classical statistical model and favors the HK statistics. An additional plot for the 10-s precipitation time series in Iowa, USA, displayed in Figure 8, is depicted in Figure 11. Here, the Hurst coefficient is very high, H ¼ 0.96. The difference between the empirical points based on classical statistics on the

one hand and the HK statistics on the other hand is quite distinctive. Apparently, the classical model is completely inappropriate for the rainfall process. The HK stochastic processes can be readily extended in a 2D setting (or even a multidimensional one). The 2D version of Equation (12) is

s ðkÞ ¼

s k 22H

ð17Þ

This can be obtained by substituting k2 for k in Equation (12). Equations (15) and (16) still hold, provided that n is the

2.7

Empirical, classical estimate Empirical, HK estimate Classical statistical model

Log(standard deviation in mm)

2.6

HK model 2.5

2.4

2.3

2.2

2.1 0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

Log(scale in years)

2.5 Empirical, classical estimate Empirical, HK estimate

Log(standard deviation in mm)

2.4

Classical statistical model HK model

2.3

2.2

2.1

2

1.9

1.8 0

0.2

0.4

0.6

0.8

1

1.2

1.4

Log(scale in years) Figure 10 Climacogram of the annual precipitation time series at: (upper) Seoul, Korea and (lower) Charleston City, USA, which are shown in Figure 7; the estimated Hurst coefficients are 0.76 and 0.74, respectively.

Precipitation

41

0.9

Log(standard deviation in mm h−1)

0.85 0.8 0.75 0.7 0.65 0.6 0.55 Empirical, classical estimate 0.5

Empirical, HK estimate Classical statistical model

0.45

HK model 0.4 0

0.5

1

1.5

2

2.5

3

Log(scale in 10-s intervals) Figure 11 Climacogram of the 10-s precipitation time series in Iowa, USA, displayed in Figure 8; the estimated Hurst coefficient is 0.96.

1.5

Log(standard deviation in mm h−1)

1.4

1.3

1.2

Empirical, classical estimate

1.1

Empirical, HK estimate Classical statistical model HK model

1 0

0.2

0.4

0.6

0.8

1

Log(scale in 0.25° of latitute and longitude) Figure 12 Climacogram of the spatial daily rainfall over the area 91 N–51 S and 78–921 E (Indian Ocean south-east of Sri Lanka) on 9 January 2006, as shown in the lower-left panel of Figure 5; the estimated Hurst coefficient is 0.94.

number of points, which is inversely proportional to k2. Figure 12 demonstrates this behavior by means of a climacogram for the spatial daily rainfall over the area 91 N–51 S and 78–921 E (Indian ocean south-east of Sri Lanka) on 9 January 2006, displayed in the lower-left panel of Figure 5. Here, the estimated Hurst coefficient is again very high,

H ¼ 0.94. As in all previous cases, the classical model is completely inappropriate, while the HK model seems reasonable for scales X4, which correspond to a resolution of 11 11 and beyond. Thus, the evidence presented using several examples of different spatial and temporal scales indicates that HK

42

Precipitation

dynamics is consistent with the nature of rainfall. These dynamics appear as scaling behavior, either in time or in space, which is either full, applicable to the entire range of scales, or asymptotic, applicable to large scales. Both these scaling behaviors are manifested as power laws of standard deviation versus temporal or spatial scale and of autocorrelation versus lag. There exists another type of scaling behavior in precipitation, the scaling in state, which is sometimes confused with the other two scaling behaviors, but is fundamentally different. Scaling in state is a property of the marginal distribution function of rainfall (it has no relation to the dependence structure of the process unlike other types of scaling) and is expressed by power laws of the tails of (1) the probability density function f(x), (2) the survival function (or exceedance probability) F  ðxÞ :¼ Pfx4 xg ¼ 1  FðxÞ, and (3) the return period T ¼ d/F*(x) where d is the length of the timescale examined. These scaling properties are expressed as

xp T k ;

F  ðxÞp x 1=k ;

f ðxÞp x 11=k

ð18Þ

and are equivalent to each other. All these are asymptotic, that is, they hold only for large values of x or, in other words, for the distribution tails. Such tails are known by several names, such as long, heavy, strong, power-type, overexponential, algebraic, or Pareto tails. The latter name comes from the Pareto distribution, which in its simplest form is given in Equation (18), although its generalized form is applicable to rainfall (see Section 2.02.5.2). As this is an asymptotic behavior, long records are needed to observe it. Figure 13 shows a logarithmic plot of the empirical distribution (expressed in terms of return period T) of a large data set of daily rainfall. This data set was created using records of 168 stations worldwide, each of which contained data of 100 years or more (Koutsoyiannis, 2004b). For each station with n years of record, n annual maximum values of daily rainfall were extracted. These values

were standardized by their mean and merged into one sample of length 17 922 station-years. From the theoretical distributions, also plotted in Figure 13, it is observed that the Pareto distribution (whose right tail appears as a straight line in the logarithmic plot; see Section 2.02.5.2) with k ¼ 0.15 provides the best fit, thus confirming the applicability of asymptotic scaling in state and the inappropriateness of the exponential-type tail. This has severe consequences, particularly in hydrological design, as distributions with exponential tails have been most common in hydrological practice, whereas it is apparent that the power-type tails are more consistent with reality. As shown in Figure 13, the difference between the two types can be substantial. Koutsoyiannis (2005a, 2005b) produced the aforesaid different types of scaling from the principle of maximum entropy. As entropy is a measure of uncertainty, the applicability of the principle of maximum entropy and its consistence with observed natural behaviors characterizing the precipitation process underscores the dominance of uncertainty in precipitation.

2.02.2 Physical and Meteorological Framework Atmospheric air is a heterogeneous mixture of gases, also containing suspended particles in liquid and solid phase. The most abundant gases are nitrogen (N2) and oxygen (O2) that account for about 78% and 21%, respectively, by volume of the atmospheric permanent gases, followed by argon (Ar) and traces of other noble gases. Their concentrations are almost constant worldwide and up to an altitude of about 90 km. Water vapor (H2O) appears in relatively low concentrations, which are highly variable. However, water vapor is very important for energy exchange on Earth (it accounts for 65% of the radiative transfer of energy in the atmosphere; Hemond

10 Empirical Pareto/least squares Pareto/L-moments Exponential/L-moments

Rescaled rainfall depth

5

2

1

0.5 1

2

5 10

100

1000

10 000

100 000

Return period Figure 13 Logarithmic plot of rescaled daily rainfall depth vs. return period: empirical estimates from a unified sample over threshold, formed using rainfall data from 168 stations worldwide (17 922 station-years). The unified sample was rescaled by the mean of each station, and fitted using a Pareto and an exponential distribution model. Adapted from Koutsoyiannis D (2004b) Statistics of extremes and estimation of extreme rainfall: 2. Empirical investigation of long rainfall records. Hydrological Sciences Journal 49(4): 591–610.

Precipitation

and Fechner-Levy, 2000), as well as for mass transfer processes in the hydrological cycle. Under certain conditions (i.e., pressure and temperature), water vapor can transform into droplets or ice crystals with subsequent release of latent heat (see Sections 2.02.2.1 and 2.02.2.2). More generally, the water-vapor content of atmospheric air affects its density, and it is of central importance in atmospheric thermodynamics (Section 2.02.2.1). The varying content and importance of water vapor in precipitation processes and thermodynamics has led to the study of atmospheric air as a mixture of two (ideal) gases: dry air and water vapor. This mixture is usually referred to as ‘moist air’ and has thermodynamic properties determined by its constituents (e.g., Rogers and Yau, 1996; Cotton and Anthes, 1989). The particles of solid and liquid material suspended in air are called ‘aerosols’. Common examples of aerosols are water droplets and ice crystals (called ‘hydrometeors’), smoke, sea salt (NaCl), dust, and pollen. The size distribution of solid aerosols depends strongly on their location. For example, the size spectrum of aerosols over land is narrow with high concentrations of small particles (e.g., kaolinite, dust, and pollen), whereas the size spectrum of aerosols over sea is wider with small concentrations of larger particles (e.g., sea salt; Ryan et al., 1972). Existence of aerosols in the atmosphere is of major importance, since a select group of aerosols called ‘hydroscopic nuclei’, is crucial for the nucleation of liquid water and initiation of rain (e.g., Brock, 1972, and Section 2.02.2.2). When moist air is cooled (i.e., below its dew point; see Section 2.02.2.1), an amount of water vapor condenses and a cloud forms, but precipitation may or may not occur. Initiation of rain requires the formation of hydrometeors (i.e., water droplets and ice crystals) of precipitable size (e.g., Gunn and Kinzer, 1949; Twomey, 1964, 1966; Brock, 1972). Formation and growth of these particles are governed by processes that take place at scales comparable to their size (micrometers to millimeters). The latter processes form the core of cloud microphysics, whereas large-scale processes, related to thermodynamics of moist air and motion of air masses, form the core of cloud dynamics. Importantly,

precipitation is the combined effect of both large- and microscale processes, and both processes are equally important and necessary for precipitation to occur.

2.02.2.1 Basics of Moist Air Thermodynamics In a parcel of moist air at temperature T with volume V and mass M ¼ Md þ Mv, with the two components denoting mass of dry air and water vapor, respectively, the density is r ¼ M/V and the concentration of water vapor, known as specific humidity, is q: ¼ Mv/M. The quantity r ¼ Mv/Md: ¼ q/(1  q) is usually referred to as the mixing ratio. The total pressure of the moist air in the parcel, p (the atmospheric pressure), equals the sum of the partial pressures of dry air pd and that of water vapor e (i.e., p ¼ pd þ e). Specific humidity and vapor pressure are interrelated through

q¼

ee p  ð1  eÞe

ð19Þ

where e ¼ 0.622 is the ratio of molar masses of water vapor and dry air. Air cannot hold an arbitrarily high quantity of vapor. Rather, there is an upper limit of the vapor pressure e*, called the saturation vapor pressure, which depends on the temperature T and is given by the Clausius–Clapeyron equation. A useful approximation to this equation is

e  ðTÞ ¼ 6:11 exp



 17:67T T þ 237:3

Saturation specific humidity, q* (g kg−1)

10

1

−30

−20

−10

0

10

Temperature, T (°C) Figure 14 Saturation specific humidity as a function of air temperature.

ð20Þ

where e* is in hPa and T is in 1C. Consequently, from Equations (19) and (20) we can calculate the saturation specific humidity q*, which is a function of T, and expresses the water vapor holding capacity of air. As shown in Figure 14, this capacity changes drastically, almost exponentially, with temperature, so that a change of temperature from  40 to 40 1C increases this capacity by 2.5 orders of magnitude. The ratio of the actual to saturation vapor pressure, that is, e/e* ¼ : U, called the relative humidity, is normally lesser than 1. When an air parcel cools, while e remains constant, e*

100

0.1 −40

43

20

30

40

44

Precipitation

decreases and hence U increases, up to the saturation value 1 or 100%. The temperature Td at which saturation occurs is called the dew point temperature and is calculated using Equation (20) by setting e*(Td) ¼ e. Therefore, cooling of the air parcel below the dew point temperature results in condensation, or transformation of the excess water vapor into liquid water in the form of droplets. During this change of phase, the relative humidity remains 100%. Condensation releases heat at a fairly constant rate (LE2.5 MJ kg1); this rate equals that of evaporation of water at a constant temperature and is thus called latent heat. For an air parcel to ascend and expand spontaneously, so that condensation and cloud formation can occur, the ambient (atmospheric) temperature gradient g : ¼  dT/dz, where z denotes altitude, also known as lapse rate, must be high (otherwise, an uplifted air parcel will sink again). While the parcel ascends and expands adiabatically (i.e., in a way that no heat transfer takes place between the air parcel and its ambient air), its own lapse rate is gd ¼ 9.8 1C km1 if the expansion is dry adiabatic (i.e., if it takes place without condensation of water vapor) and somewhat smaller, g*, if the expansion is moist adiabatic (i.e., if the temperature has fallen below dew point, so that some of the water vapor in the parcel condenses to liquid form). The gradient g* is not constant but varies with temperature T and air pressure p so that g* ¼ 4 1C km1 for T ¼ 25 1C and p ¼ 1000 hPa, whereas g* ¼ 9 1C km1 for T ¼  25 1C and p ¼ 1000 hPa; an average value is g* ¼ 6.5 1C km1 (Koutsoyiannis, 2000b; see also Wallace and Hobbs, 1977). When the ambient lapse rate g is smaller than g*, the atmosphere is stable, and no spontaneous lift occurs and no clouds are formed. When g4gd, the atmosphere is unstable and favors air lift and formation of clouds. The case g*ogogd is known as conditional instability and it serves as an important mechanism for mesoscale precipitation processes (see Sections 2.02.2.4 and 2.02.2.5).

2.02.2.2 Formation and Growth of Precipitation Particles The Clausius–Clapeyron equation describes the equilibrium condition of a thermodynamic system consisting of bulk water and vapor. A state out of the equilibrium, in which e4e* (U41) is possible, but is thermodynamically unstable, and is called supersaturation. Detailed study of the transition of water vapor to liquid or ice at or above saturation is associated with certain free-energy barriers. An example of such an energy barrier is the dynamic energy associated with the surface tension, s, of a water droplet. For a spherical droplet, s is proportional to the pressure of water within the droplet p and inversely proportional to its radius r (i.e., s ¼ p/2r). This means that a high vapor pressure is needed for a very small droplet to be maintained and not evaporate. In essence, the free-energy barrier of surface tension makes droplet formation solely by condensation of water vapor (a process usually referred to as ‘homogeneous nucleation’), almost impossible in nature. However, if the surface-tension barrier is bypassed, common supersaturations of the order of 1–2% (i.e., U ¼ 1.01–1.02) are sufficient for water vapor to diffuse toward the surface of the droplet. The rate of diffusional growth is proportional to the supersaturation U  1 of the ambient air,

and inversely proportional to the radius r of the droplet: that is, dr/dtp(U  1)/r (Mason, 1971; Rogers and Yau, 1996). While homogeneous nucleation requires large supersaturations, formation of droplets is drastically facilitated by particulated matter of the size of micrometers or lower, the aerosols, some of which, called condensation nuclei, are hydrophilic and serve as centers for droplet condensation (Brock, 1972; Slinn, 1975; Hobbs et al., 1985). This process is usually referred to as ‘heterogeneous nucleation’ and it is almost exclusively the process that governs water vapor condensation in the atmosphere (Houze, 1993). When the temperature in the cloud drops below the freezing point, water droplets are said to be supercooled, and they may or may not freeze. For pure water droplets, homogeneous freezing does not occur until the temperature drops below  40 1C (Rogers and Yau, 1996). However, the presence of certain condensation nuclei, called ice nuclei, may allow freezing of water droplets at temperatures a few degrees below 0 1C. These nuclei are particles of the size of micrometers, or lower, which form strong bonds with water and closely match the crystallic structure of ice. Different particles serve as condensation nuclei at different subfreezing temperatures. For example, silver iodide (AgI) serves as an ice nucleator at  4 1C and kaolinite at  9 1C (e.g., Houghton, 1985). Evidently, a cloud is an assembly of tiny droplets with usually met concentrations of several hundreds per cubic centimeter, and radii of several micrometers. This structure is very stable and the only dominant process is vapor diffusion, which accounts for the size-growth evolution of the whole droplet population (Telford and Chai, 1980; Telford and Wagner, 1981). Precipitation develops when the cloud population becomes unstable and some droplets grow faster relative to others. In general, two main mechanisms account for the cloud microstructure becoming unstable. The first mechanism is the collision and coalescence (i.e., sticking) of larger (and fastermoving) collector drops with smaller (and slower-moving) collected droplets. This mechanism is particularly important for precipitation development in warm clouds (i.e., at temperatures in excess of 0 1C; see, e.g., Houze, 1993) and, for a long time, it has formed an active research area in cloud and precipitation physics (e.g., Langmuir, 1948; Bowen, 1950; Telford, 1955; Scott, 1968, 1972; Long, 1971; Drake, 1972a, 1972b; Gillespie, 1972, 1975; Robertson, 1974; Berry and Reinhardt, 1974a, 1974b; Vohl et al., 1999; Pinsky et al., 1999, 2000; Pinsky and Khain, 2004; and review in Testik and Barros (2007)). Its significance for precipitation processes depends considerably on the droplet-size spectra, with larger effectiveness for wider spectra with small concentrations of larger particles (Berry and Reinhardt, 1974a, 1974b). The second mechanism is related to interaction between water droplets and ice crystals, and is limited to clouds with tops that extend to subfreezing temperatures (i.e., cold clouds). In particular, when an ice crystal develops in the presence of a large number of supercooled droplets, the situation becomes immediately unstable and the ice crystal grows due to diffusion of water vapor from the droplets toward the crystal. This is due to the fact that the equilibrium vapor pressure over ice is less than that over water at the same subfreezing temperature. Thus, the ice crystal grows by

Precipitation

diffusion of water vapor and the supercooled droplets evaporate to compensate for this. The transfer rate of water vapor depends on the difference between the equilibrium vapor pressure of water and ice, a quantity that becomes sufficiently large at about  15 1C (Uijlenhoet, 2008). The latter process is called the Bergeron–Findeisen mechanism, named after the scientists who first studied it (Bergeron, 1935; Findeisen, 1938). Once the ice crystals have grown by vapor diffusion to sizes sufficiently large for gravitational settling to dominate, they start falling and colliding with their ambient droplets and ice crystals, a process usually referred to as ‘accretional growth’. In the first case (i.e., when ice crystals collide with droplets), graupel or hail may form, whereas in the second case, snowflakes are likely to form. As the frozen particles fall, it is possible to enter layers with temperatures higher than 0 1C and start melting. If the particles have relatively small terminal velocities (or equivalently small size; see Section 2.02.2.3), they may reach the ground as raindrops indistinguishable from those formed by coalescence. Alternatively, in cold weather, or when large hailstones are formed, the precipitation particles may reach the ground unmelted. Additional discussion on the mechanisms of formation and growth of precipitation particles, and the potential human intervention on the mechanisms by technological means are discussed in Chapter 4.05 Abstraction of Atmospheric Humidity.

2.02.2.3 Properties of Precipitation Particles 2.02.2.3.1 Terminal velocity The terminal velocity UX(D) of a precipitable particle of type X ¼ R (rain), H (hail), S (snow), and effective diameter D is the maximum velocity this particle may develop under gravitational settling relative to its ambient air. In theory, UX(D) can be obtained by balancing the weight of the particle with the sum of the static and dynamic buoyancy (i.e., drag forces) on the particle. For a rigid spherical raindrop, one obtains pffiffiffiffi UR ðDÞp D (e.g., Rogers and Yau, 1996). Theoretical calculation of UX(D) becomes more complicated when the dynamical characteristics of the falling particles depend on their linear size D and the ambient temperature T. For example, droplets with diameters D smaller than about 0.35 mm are approximately spherical, drops with diameters in the range 0.35–1 mm tend to deform by the aerodynamic shear receiving a more elliptical shape, whereas larger drops frequently break down into smaller droplets due to excessive elongation or surface vibrations (e.g., Testik and Barros, 2007; Uijlenhoet, 2008). Moreover, the crystallic structure, shape, size, and, hence, the aerodynamic properties of snowflakes depend on the ambient temperature T (Fletcher, 1962; Locatelli and Hobbs, 1974; Houghton, 1985; Rogers and Yau, 1996). In the absence of exact theoretical solutions for the terminal velocity UX(D) of precipitation particles under complex atmospheric conditions, several empirical formulae have been developed (e.g., Gunn and Kinzer, 1949; Liu and Orville, 1969; Wisner et al., 1972; Locatelli and Hobbs, 1974; Atlas and Ulbrich, 1977; Lin et al., 1983). According to Liu and

45

Orville (1969), who performed a least squares analysis of Gunn and Kinzer’s (1949) data, the terminal velocity of raindrops of diameter D can be approximated by a power-law type relationship:

UR ðDÞ ¼ a Db

ð21Þ

where a ¼ 2115 cm1b s1 and b ¼ 0.8 are empirical constants. For raindrops with diameters in the range 0.5pDp5 mm, Atlas and Ulbrich (1977 (see also Uijlenhoet, 2008) suggest the use of Equation (21) with parameters a ¼ 1767 cm1b s1 and b ¼ 0.67. For hail, Wisner et al. (1972) suggest

UH ðDÞ ¼ D1=2



4grH 3CD r

1=2 ð22Þ

where g ¼ 9.81 m s2 is the acceleration of gravity, rE1.2 kg m3 is the density of air, rH ¼ 800–900 kg m3 is the density of the hailstone, and CD ¼ 0.6 is a drag coefficient. For graupel-like snow of hexagonal type, Locatelli and Hobbs (1974) suggest:

US ðDÞ ¼ c Dd

ð23Þ

where c ¼ 153 cm1d s1 and d ¼ 0.25 are empirical constants that, in general, depend on the shape of the snowflakes (e.g., Stoelinga et al., 2005). UX(D) relationships other than power laws have also been suggested (e.g., Beard (1976) and review by Testik and Barros (2007)). However, the power-law form in Equations (21)–(23) is the only functional form that is consistent with the powerlaw relations between the radar reflectivity factor Z (see Section 2.02.3.2) and the rainfall intensity i (Uijlenhoet, 1999, 2008).

2.02.2.3.2 Size distribution A commonly used parametrization for the size distributions of precipitation particles is that introduced by Marshall and Palmer (1948). According to this parametrization, precipitation particles have exponential size distributions of the type

nX ðDÞ ¼ n0X expðbX DÞ;

X ¼ R; H; S

ð24Þ

where the subscript X denotes the type of the particle: rain (R), hail (H), or snow (S); D is the effective diameter of the particle; bX is a distribution scale parameter with units of (length1) (see below); and n0X is an intercept parameter that depends on the type of the particle with units of (length4): that is, number of particles per unit diameter and per unit volume of air (see below). To determine the parameters n0R and bR in Equation (24) for rainfall, Marshall and Palmer (1948) used observations from summer storms in Canada. The study reported a constant value of the intercept parameter n0R ¼ 8 102 cm4, whereas the scale parameter bR was found to vary with the rainfall intensity i at ground level as: bR ¼ 41 i0.21 cm1, where i is in millimeters per hour. Clearly, the mean raindrop size 1/bR increases with increasing rainfall intensity i.

46

Precipitation

Gunn and Marshall (1958) used snowfall observations from Canada to determine the parameters n0S and bS for snow. The study concluded that both n0S and bS depend on the precipitation rate as:

n0S ¼ 0:038 i0:87 cm4 ;

bS ¼ 25:5 i0:48 cm1

2.02.2.4 Clouds and Precipitation Types ð25Þ

where i is the water equivalent (in millimeters per hour) of the accumulated snow at ground level. Similar to the mean raindrop size, the mean snowflake size 1/bS increases with increasing i. A modification to the distribution model of Gunn and Marshall (1958) has been proposed by Houze et al. (1979) and Ryan (1996). According to these authors, the intercept parameter for snow, n0S, is better approximated as a decreasing function of the temperature T of the ambient air. The latter is responsible for the properties and structures of ice crystals (see Section 2.02.2.2). Federer and Waldvogel (1975) used observations from a multicell hailstorm in Switzerland to determine the parameters n0H and bH for hail. The study showed pronounced variability of the intercept parameter n0H ¼ 15 106 to 5.2 104 cm4, moderate variability of the scale parameter bH ¼ 3.3–6.4 cm1, and concluded showing an exponential mean size distribution for hailstones with constant parameters: n0HE1.2 104 cm4 and bHE4.2 cm1. Alternative models, where the size distributions of precipitation particles are taken to be either gamma or lognormal, have also been suggested (e.g., Ulbrich, 1983; Feingold and Levin, 1986; Joss and Waldvogel, 1990). However, the exponential distribution model introduced by Marshall and Palmer (1948) has been empirically validated by a number of studies (see, e.g., Kessler, 1969; Federer and Waldvogel, 1975; Joss and Gori, 1978; Houze et al., 1979; Ryan, 1996; Ulbrich and Atlas, 1998; Hong et al., 2004), and has found the widest application by being used in the cloud-resolving schemes of many state-of-the-art NWP models (e.g., Cotton et al., 1994; Grell et al., 1995; Reisner et al., 1998; Thompson et al., 2004; Skamarock et al., 2005). A more general formulation for the size distribution of precipitation particles, which includes the exponential model of Marshall and Palmer (1948), and the gamma and lognormal models as special cases, was suggested by SempereTorres et al. (1994, 1998). According to their formulation, the size distribution of precipitation particles can be parametrized as

nX ðDÞp if gðDX =iz Þ;

X ¼ R; H; S

ð26Þ

where f and z are constant exponents, i is the precipitation rate, and g(x) is a scalar function with parameter vector a. For a certain form of g, the functional dependence of the parameters f, z, and a is obtained by satisfying the equation for the theoretical precipitation rate originating from particles with size distribution nX(D) (Sempere-Torres et al., 1994, 1998; Uijlenhoet, 2008):

i¼

p 6

ZN 0

nX ðDÞUX ðDÞD3 dD;

X ¼ R; H; S

where UX(D) is given by Equations (21)–(23). Note, however, that the units of nX depend on those used for D and i and, of course, the functional form of g(x).

ð27Þ

Clouds owe their existence to the process of condensation, which occurs in response to several dynamical processes associated with motions of air masses, such as orographic or frontal lifting (see Section 2.02.2.5), convection, and mixing. At the same time, clouds and the resulting precipitation influence the dynamical and thermodynamical processes in the atmosphere. For example, clouds affect air motions through physical processes, such as the redistribution of atmospheric water and water vapor, the release of latent heat by condensation, and the modulation of the transfer of solar and infrared (IR) radiation in the atmosphere. A cloud system is formed by a number of recognizable isolated cloud elements that are identifiable by their shape and size (e.g., Scorer and Wexler, 1963; Austin and Houze, 1972; Orlanski, 1975). On the lowest extreme, cloud systems with a scale of about 1 km or less are classified as microscale systems. On the highest extreme, atmospheric phenomena of linear extent of 1000 km and upward are classified in the synoptic scale and include the cloud systems associated with baroclinic instabilities, and extratropical cyclones (i.e., lowpressure centers). In between these two extreme scales, atmospheric phenomena with linear extent between a few kilometers and several hundred kilometers are the so-called mesoscale phenomena. These phenomena are more likely associated with atmospheric instabilities, as well as frontal and topographic lifting. Mesoscale phenomena include many types of clouds and cloud systems that are usually classified into two main categories: stratiform and convective (cumulus) cloud systems. In general, stratiform cloud systems have the shape of a flat appearing layer and produce widespread precipitation associated with large-scale ascent, produced by frontal or topographic lifting, or large-scale horizontal convergence. By contrast, convective cloud systems have large vertical development, produce localized showery precipitation, and are associated with cumulus-scale convection in unstable air. Next, we focus on the structure of these systems and the forms of precipitation they produce.

2.02.2.4.1 Cumulus cloud systems Cumulus clouds are formed by small thermals (upwardmoving air parcels heated by contact to the warm ground) where condensation occurs and they grow to extend vertically throughout the troposphere. Their vertical extent is controlled by the depth of the unstable layer, while their horizontal extent is comparable to their vertical extent. A typical linear dimension of a cumulus cloud is 3–10 km, with updraft velocities of a few meters per second (Rogers and Yau, 1996). Observations performed by Byers and Braham (1949; see also Weisman and Klemp, 1986) revealed that convective storms are formed by a number of cells, each one of which passes through a characteristic cycle of stages (Figure 15). The cumulus stage of a cell is characterized by an updraft throughout most of the cell. At this stage, which lasts approximately 10–20 min, the cell develops and expands

Precipitation

47

z ≈ 10−12 km

Rain

Rain z ≈ 6 km

New cell development Updraft

Downdrafting air

Updraft

z = 0 km 6−8 km

8−16 km

6−11 km

Cumulus stage 10−20 min

Mature stage 15−30 min

Dissipating stage ≈ 30 min

Figure 15 Stages of development of convective cells. Adapted from Weisman ML and Klemp JB (1986) Characteristics of isolated convective storms. In: Ray PS (ed.) Mesoscale Meteorology and Forecasting, ch. 15, pp. 331–358. Boston, MA: American Meteorological Society.

vertically while the air becomes saturated and hydrometeors grow due to vapor condensation and turbulent coalescence (see Section 2.02.2.2). Some ice and water particles grow large enough to fall relative to the ambient updraft and initiate a downdraft within the cell. The downdraft is initially in saturated condition, but as it moves toward the lower troposphere and mixes with subsaturated air, evaporational cooling occurs, which introduces negative buoyancy and accelerates the downdraft. This is the start of the mature stage of the cell, which lasts for approximately 15–30 min. The air of the downdraft reaches the ground, as a cold core, and changes the surface wind pattern. This change may initiate a new thermal at a neighboring location, which might grow to a new cell. The downdraft interferes with the updraft at the lower levels of the cloud and finally cuts off the updraft from its source region. At this point, the cell enters its dissipating stage. At this stage, which lasts for about 30 min, the updraft decays and consequently, the precipitation source is eliminated.

whereas thick stratus clouds (i.e., 1–2 km vertical extent) are capable of producing substantial widespread rain or snow. Although the classification of cloud systems in stratiform and convective is useful for observation purposes, it cannot be considered sharp (Harrold and Austin, 1974). Observations from radars or rain gauges show that widespread precipitation has a fine-scale structure with intense precipitation regions confined to elements with size of a few kilometers, while rainfall features of convective origin (e.g., cells) can grow and/ or cluster over a large region producing continuous precipitation similar to that of stratiform formations. In general, convective rainfall patterns are nonuniform and are associated with locally intense rainfall regions ranging in size from 3 to 10 km. The latter evolve rapidly in time and are separated by areas free of precipitation. By contrast, stratiform patterns are associated with less-pronounced small-scale structures and a wider overall extent that persists in time.

2.02.2.5 Precipitation-Generating Weather Systems 2.02.2.5.1 Fronts

2.02.2.4.2 Stratus cloud systems Stratus clouds are associated with mesoscale, or even synoptic, vertical air motions that arise from large-scale horizontal convergence and frontal or orographic lifting of moist air masses. The ascending motion of air is weak (i.e., a few tens of centimeters per second) relative to cumulus convection, but it extends over large areas and durations to produce widespread rain or snow. The lifetime of a stratus formation is of the order of days, and its size may extend over hundreds of kilometers horizontally. The ascended air masses, having the form of a flat appearing layer, remain convectively stable even after they are lifted to higher altitudes. Since atmospheric turbulence is not intense, initiation of rain is mainly dominated by the ice particle growth due to vapor deposition (the Bergeron–Findeisen mechanism; Section 2.02.2.2), when the ascended air masses are thick enough to reach subfreezing temperatures. In general, thin stratus clouds are usually nonprecipitating,

Atmospheric circulation is formed by advecting air masses with fairly uniform characteristics. Depending on their source of origin, different air masses may have different temperatures and moisture contents. For example, continental air masses are drier and their temperatures vary in a wider range relative to maritime air masses. The interface of two opposing air masses with different temperatures and moisture contents is usually referred to as a front. Along this interface, the warmer and lighter air rises above the colder and denser air. The vertical lifting causes the warmer air to cool adiabatically, the water vapor to condense, and, hence, precipitation to form. A cold front occurs when advancing cold air wedges itself under warmer air and lifts it (Figure 16(a)), whereas a warm front develops when faster-moving warm air overrides a colder and denser air mass (Figure 16(b); Koutsoyiannis and Xanthopoulos, 1999). An occluded front forms when warm air is trapped between two colder and denser air masses. An example of an occluded front is shown in Figure 16(c), where a cold front catches up a slower-moving warm front.

48

Precipitation

Upper-level winds

Cold front Warm air

Advancing cold air (a)

Warm front

Advancing warm air

Retreated cold air Cold air

(b)

Warm air Warm front Cold front Cold air Advancing cold air

Occluded front

(c)

Figure 16 Schematic illustration of different types of fronts: (a) cold front, (b) warm front, and (c) occluded front. Adapted from Koutsoyiannis D and Xanthopoulos T (1999) Engineering Hydrology, 3rd edn., p. 418. Athens: National Technical University of Athens (in Greek).

Fronts may extend over hundreds of kilometers in the horizontal direction and are associated with vertical wind speeds of the order of a few tens of centimeters per second. This range of values is in accordance with vertical motions caused by the horizontal wind convergence of synoptic-scale low-level flow. Hence, frontal precipitation is mostly stratiform with widespread rain or snow over large areas and durations. Note, however, that embedded within the areas of frontal precipitation there are mesoscale regions that exhibit cellular activity.

2.02.2.5.2 Mechanical lifting and orographic precipitation Orographic precipitation occurs when horizontally moving warm and humid air meets a barrier such as a mountain range. In this case, the barrier causes uplift of the incoming air. As the moist air moves upslope, it cools adiabatically, water vapor condenses to liquid water or ice (depending on the altitude where the dew point temperature occurs), and precipitation is likely to form (e.g., Smith, 1993; Hemond and Fechner-Levy, 2000). In general, orographic precipitation (unless combined with other mechanisms such as cyclonic activity and fronts) is narrow banded since it occurs in association with water-vapor condensation by mechanical lifting, a process that becomes

effective at a certain elevation along the topography. After surpassing the top of the mountain range, on the lee side, the air moves downward and this causes adiabatic warming, which tends to dissipate the clouds and stop the precipitation, thus producing a rain shadow.

2.02.2.5.3 Extratropical cyclones Extratropical cyclones are synoptic scale low-pressure systems that occur in the middle latitudes (i.e., pole-ward of about 301 latitude) and have length scales of the order of 500–2500 km (e.g., Hakim, 2003). They usually form when two air masses with different temperatures and moisture contents that flow in parallel, or are stationary, become coupled by a preexisting upper-level disturbance (usually a low-pressure center) near their interface. An example is the formation of extratropical cyclones along the interface of mid-latitude westerlies (i.e., winds that flow from West to East; e.g., Lutgens and Tarbuck, 1992), with the equator-ward-moving polar, and thus colder, air masses (i.e., polar easterlies). As shown in Figure 17, which refers to the Northern Hemisphere, the motion of both warm and cold air masses is caused by pressure gradients and their direction is south–north and north–south, respectively (Koutsoyiannis

Precipitation

49

er lie s

High ld

air

la r

ea

st

Co

Low

Po

Cold air

Warm

t

on d fr

`

Col

rli es

Warm air

te

r

.w es M id la t

(a)

front

m ar

ai

W

(b)

High

Cold air Low

Low

Co

ld

m

ar

W

air

Occluded front air

Warm front

(c)

(d)

m

ar

air

W

Wid stra espre a tifo rm d reg io

n

Band of precipitation

air W ar m

Co

ld

ai

r

Great Britain

France

(e) Figure 17 (a)–(d) Schematic illustration of the evolution of an extratropical cyclone at the interface of mid-latitude westerlies and the equator-wardmoving polar easterlies; and (e) extra-tropical cyclone over the British Isles on 17 January 2009: motion of air masses, fronts, and characteristic precipitation regions. (a)–(d), Adapted from Koutsoyiannis D and Xanthopoulos T (1999) Engineering Hydrology, 3rd edn., p. 418. Athens: National Technical University of Athens (in Greek); and (e) from http://www.ncdc.noaa.gov/sotc/index.php?report ¼ hazards&year ¼ 2009&month ¼ jan).

50

Precipitation

Over shooting top

Storm

motion

12−16 km Upper-level winds anvil vau lt

9 km

Rain and hail

3−6 km Mid-level winds

Forward flank downdraft

Rear flank downdraft Tornado

Gust front Figure 18 Schematic illustration of the wind circulation in a super-cell storm. Adapted from http://www.nssl.noaa.gov/primer/tornado/.

and Xanthopoulos, 1999). However, these directions are diverted to the right (in the Northern Hemisphere) by Coriolis forces. The initial disturbance formed by the shear along the interface of the two air masses (Figure 17(a)) grows as the warmer and lighter air rises above the colder air and starts rotating in an emerging spiral called the cyclone (Figure 17(b)). As the cyclone evolves, the cold front approaches the slower-moving warm front (Figure 17(c)) and then catches up with it forming an occluded front (Figure 17(d)). Finally, mixing between the two air masses causes the fronts to lose their identities and the cyclone to dissipate. The adiabatic cooling of the warm and moist air results in a widespread region of stratiform precipitation that propagates with the upper-level flow far beyond the fronts (Figure 17(e)).

2.02.2.5.4 Isolated extratropical convective storms

The super-cell storm is the most intense of all isolated convective storms. It has a lifetime of several hours, it exhibits large vertical development, and produces strong winds, heavy rainfall, or hail, and long-lived tornadoes, that is, intense vortices with diameters of the order of 100–500 m (e.g., Browning and Ludlam, 1962; Rotunno, 1986; Weisman and Klemp, 1986; Bluestein, 2003), where the updrafts and downdrafts are displaced horizontally and interact mutually to sustain a long-lived circulation (Figure 18). The updraft enters at low levels and ascends in a region called the vault, which might penetrate into the stratosphere. Super-cell storms usually evolve from multicell formations when the magnitude of the vertical wind shear, defined as the difference between the density-weighted mean wind over the lowest 6 km and a representative surface layer wind (e.g., 500 m mean wind), suffices to produce a long-lived rotating updraft that mutually interacts with the downdraft (e.g., Weisman and Klemp, 1982, 1984, 1986).

A short-lived single-cell is the simplest storm of convective origin. Single cells have horizontal cross sections of the order of 10–100 km2 and move with the mean environmental flow over the lowest 5–7 km of the troposphere. The stages of development of a single-cell storm were discussed in Section 2.02.2.4. The multicell storm is a cluster of short-lived single cells with cold outflows (i.e., downdrafts) that combine to form a large gust front (Weisman and Klemp, 1986). The convergence along the leading edge of the front triggers new updraft development and subsequent formation of new cells. Because of the new cell development, multicell storms may last several days and span over large areas with linear extents of hundreds of kilometers.

Intense rainfall events are usually organized in lines (i.e., squall lines) and bands (i.e., rainbands) with characteristic scales of hundreds of kilometers. According to Hane (1986), rainbands are sufficiently elongated rainfall areas that are nonconvective or weakly convective, and squall lines include all linear convective structures stronger than rainbands. These large-scale features are considered to be manifestations of the large mesoscale horizontal circulation, in association with spatial fluctuations of the surface temperature and moisture content of atmospheric air masses.

2.02.2.5.5 Extratropical squall lines and rainbands

Precipitation

The conditions for squall line formation are (1) a convectively unstable near-surface environment (i.e., moist and warm near-surface air with relatively cold air aloft) to maintain the development of convective cells, (2) a layer of dry air directly above the near-surface moist air to enhance development of an intense and wide cold downflow by evaporative cooling (i.e., the dry middle-level air causes precipitation particles to evaporate and a negatively buoyant cold front to form), and (3) a triggering mechanism for release of the convective instability (e.g., frontal or orographic lifting). Once the squall line has formed, it feeds itself through convergence along the cold gust front. This convergence produces strong ascent and forms new cells ahead of the storm. Rainbands in extratropical regions occur primarily in association with well-organized extratropical cyclones (Hane, 1986). In this case, precipitation is maintained by the ascent resulting from the warm advection of the advancing cyclone, with subsequent formation of a widespread region of stratiform precipitation (Section 2.02.2.5.3). Extratropical rainbands can also be formed in synoptic-scale environments other than those associated with cyclonic circulation. An example is the environment associated with the development of symmetric instabilities (e.g., Bennetts and Sharp, 1982; Seltzer et al., 1985).

2.02.2.5.6 Monsoons The term monsoon generally applies to climates that exhibit long, distinct, and remarkably regular rainy and dry periods associated with the spatial distribution of solar heating during summer and winter. According to a definition proposed by Ramage (1971), a monsoon climate is characterized by (1) prevailing wind directions that shift by at least 1201 between January and July, (2) prevailing wind direction that persists at least 40% of the time in January and July, (3) mean wind speeds that exceed 3 m s1 in either January or July, and (4) fewer than one cyclone–anticyclone alternation every 2 years in either January or July in a 51 latitude–longitude rectangle. In essence, Ramage’s (1971) criteria exclude most extratropical regions with prevailing synoptic-scale cyclonic and anticyclonic circulations and, in addition, require the mean wind direction to be driven and sustained exclusively by the seasonally varying temperature contrast between continental and oceanic masses. Under these constraints, only India, SouthEastern Asia, Northern Australia, and West and central Africa have monsoon climates (Slingo, 2003). For example, in India, about 80% of the mean annual rainfall accumulation (about 2 m) occurs during the months of June, July, and August (Smith, 1993). The main driving mechanism for monsoons is the temperature contrast between continental and oceanic masses due to the seasonal cycle of solar heating. More precisely, the lower thermal inertia of continental masses relative to oceans causes the former to heat up more rapidly during spring and summer by the solar radiation. This results in a sharp temperature gradient, which causes a humid flow of oceanic near-surface air to move toward the land (something similar to a massive sea breeze). As it reaches the land, the humid air warms up and rises, water vapor condenses to liquid water, and rain falls. A similar process occurs during winter, when the continental

51

air masses cool up more rapidly than the surrounding ocean water, with subsequent formation of a cold and dry massive low-level flux toward the ocean. An important factor that determines the intensity of monsoon rainfall is the geographical orientation of continents and oceans relative to the equator (Slingo, 2003). For example, the north–south orientation of the South-Eastern Asian and Northern Australian monsoon system allows the dry outflow from the winter continent to warm up and load moisture from the ocean, flow across the equator toward the summer hemisphere, and, eventually, feed the monsoon rains over the summer continent. This is also the reason why the largest rainfall accumulations for durations larger than 24 h are associated with the Asian–Australian monsoon system (Smith, 1993).

2.02.2.5.7 Tropical cyclones Tropical cyclones form a particular class of synoptic-scale lowpressure rotating systems that develop over tropical or subtropical waters (Anthes, 1982; Landsea, 2000). These systems have linear extent of the order of 300–500 km and are characterized by well-organized convection and cyclonic (counterclockwise in the Northern Hemisphere) surface wind circulation around a relatively calm low-pressure region, called the eye of the storm (Figures 19 and 20). Tropical cyclones with sustained wind speeds in the range 17–32 m s1 are called tropical storms whereas stronger tropical cyclones are usually referred to as hurricanes (i.e., when observed in the North Atlantic Ocean, in the Northeast Pacific Ocean east of the dateline, and in the South Pacific Ocean east of 1601 E) or typhoons (i.e., when observed in the Northwest Pacific Ocean west of the dateline). Note, however, that extreme rainfall accumulations for durations of the order of a day, or higher, are usually produced by moderate or even low-intensity tropical cyclones (Langousis and Veneziano, 2009b). An example is the tropical storm Allison in 2001, which looped over the Houston area causing rainfall accumulations in excess of 850 mm. According to the US National Oceanic and Atmospheric Administration (NOAA; Stewart, 2002), Allison (2001) ranks as the costliest and deadliest tropical storm in the history of the US with 41 people killed, 27 of who were drowned, and more than $6.4 billion (2007 USD) in damages. The genesis and development of tropical cyclones require the following conditions to be maintained (e.g., Gray, 1968, 1979): (1) warm ocean waters (surface temperature T427 1C); (2) a conditionally unstable atmosphere where the air temperature decreases fast with height; (3) a relatively moist midtroposphere to allow the development of widespread thunderstorm activity; (4) a minimum distance of about 500 km from the equator in order for the Coriolis force to be sufficiently large to maintain cyclonic circulation; (5) a nearsurface disturbance with sufficient vorticity and low-level convergence to trigger and maintain cyclonic motion; and (6) low magnitude of vertical wind shear (less than 10 m s1), defined as the difference between the 200- and 850-hPa horizontal wind velocities in the annular region between 200 and 800 km from the tropical cyclone center (Chen et al., 2006). The latter condition is important for the maintenance of the deep convection around the center of the cyclone.

52

Precipitation Tropopause, Z ≈ 15 km Main vortex top Z ≈ 10 km

Divergence region

Outflow Eye

Cloud

Rain Main vortex

Vertical wind

Boundary layer top, H ≈ 1−2 km Surface boundary, Z=0

Inflow

Boundary layer

R ≈ 150−200 km

R = 0 R ≈ 15−40 km Figure 19 Schematic representation of the structure of a mature hurricane.

30

d

an

b ain r-r

Storm motion

te

Ou

29

≥18 16

Inner-rainbands

14

Latitude (deg)

28

12 10 27

8 6 4

26

25

Eyewall

2 0 (mm h−1)

Out

er-r ainb

and Katrina (2005)

24 −92

−91

−90

−89

−88

−87

−86

Longitude (deg) Figure 20 TRMM microwave imager (TMI) rainfall retrievals for Hurricane Katrina on 28 August (2005) at 21:00 UTC (frame 44373): different types of rainbands and their locations relative to the center of the storm.

At a first approximation, a tropical cyclone can be seen as a heat engine fueled by the buoyant motion of warm and saturated (hence convectively highly unstable) air masses that lie directly above the warm tropical and subtropical ocean waters (e.g., Emanuel, 1986, 1989; Renno and Ingersoll, 1996; Marks, 2003). By contrast, extratropical cyclones obtain their energy

from the horizontal temperature gradients in the atmosphere (Section 2.02.2.5.3). During its mature stage, a tropical cyclone includes four distinct flow regions (Yanai, 1964; Smith, 1968; Frank, 1977; Willoughby, 1990; Smith, 2000), as depicted in Figure 19:

Precipitation

1. Away from the surface boundary (in the altitude range from 2–3 km to about 10 km), frictional stresses are negligible and the horizontal winds are in approximate gradient balance (e.g., La Seur and Hawkins, 1963; Hawkins and Rubsam, 1968; Holland, 1980; Willoughby, 1990, 1991; Vickery et al., 2000). In this region, usually referred to as the main vortex, the radial inflow is negligible, whereas the tangential flow is maintained by the balance between the inward-directed pressure gradient force and the sum of the outward-directed centrifugal and Coriolis forces. 2. Within the boundary layer (in the altitude range below 1– 2 km), frictional stresses decelerate the tangential flow, reduce the magnitude of the Coriolis and centrifugal forces, and result in an inward net force that drives low-level convergence. Calculations performed by Smith (1968, 2003), Kepert (2001), Kepert and Wang (2001), and Langousis et al. (2008) show that the radial inflow in the boundary layer turns upward before it reaches the tropical cyclone center causing vertical fluxes of moisture. Langousis and Veneziano (2009a) showed that these fluxes can be used to obtain accurate estimates for the large-scale mean rainfall intensity field in tropical cyclones as a function of the tropical cyclone characteristics. 3. At altitudes in excess of about 10 km, the curved isobars, which are responsible for the tropical cyclone formation and maintenance, start to flatten. As a consequence, the inward-directed pressure gradient force that maintains the cyclonic circulation decreases with increasing height leading to an outward-directed net force that drives high-level divergence. 4. Finally, there is a core flow region, called the eye of the tropical cyclone, with diameters of the order of 15–40 km. This region is free of cloud with light tangential winds and a downflow close to the axis. The condensation of water vapor caused by the ascending motion of humid near-surface air leads to the formation of cloud systems. These systems, which are usually precipitating, are organized around the cyclone center into long quasi-circular formations usually referred to as rainbands. Despite variations of rainband characteristics from one storm to another, and during the evolution of a single storm (e.g., Miller, 1958; Barnes et al., 1983; Marks, 1985; Molinari et al., 1999), a number of studies (Willoughby et al., 1984; Powell, 1990; Molinari et al., 1994, 1999, among others) have shown that rainbands, depending on their location relative to the storm center, share similar structural characteristics and can be organized into three distinct classes: eyewall, inner-rainbands, and outer-rainbands (Figure 20): 1. The eyewall is a well-developed convective band that surrounds the eye of the tropical cyclone. This band has a width of approximately 10–15 km with upward-directed quasi-steady velocities in the range of 0.5–3 m s1 or more, with the larger values being associated with more intense systems. The quasi-steady updrafts mostly reflect the radial convergence of horizontal fluxes, which become maximum close to the eye of the tropical cyclone (Smith, 1968; Shapiro, 1983; Kepert, 2001). The eyewall almost always

53

has the highest cloud tops (Jorgensen, 1984a), contains the largest annular mean rainfall intensity (Marks, 1985; Houze et al., 1992), and exhibits weak cellular structure as evidenced by radar observations (e.g., Jorgensen, 1984b; Marks, 1985). 2. The inner-rainbands (Molinari et al., 1994, 1999) are a group of spiral bands located outside the eyewall at radial distances smaller than approximately 120 km, and are also referred to as a stationary band complex (Willoughby et al., 1984). This group moves slowly, if at all, and maintains a rather fixed position relative to the vortex. Rainfall inside the inner-rainband region is mostly stratiform, with active convection covering 5–10% of the total rainfall area and contributing 40–50% of the total rainfall volume (e.g., Marks, 1985; Marks and Houze, 1987; Marks et al., 1992). 3. Outer-rainbands typically occur at radial distances larger than approximately 150 km from the tropical cyclone center (e.g., Powell, 1990; Molinari et al., 1994). They develop by the increased convergence at the boundary of the vortex envelope, where the convectively unstable environmental air flows around the storm and gives rise to formation of convective cells (e.g., Beer and Giannini, 1980; Ooyama, 1982; Molinari et al., 1994). Consequently, outerrainbands have more cellular structure than inner ones, which develop in a less-unstable atmosphere.

2.02.3 Precipitation Observation and Measurement 2.02.3.1 Point Measurement of Precipitation 2.02.3.1.1 Measuring devices The measurement of precipitation at a point is as easy as placing a bucket at the point of observation and periodically measuring the quantity of water it collects. The collected volume divided by the area of the opening is the precipitation depth. Due to this simplicity, such gauges have been used systematically since many centuries, and must have been discovered independently in different times, perhaps even in the antiquity, and in different places in the world, such as in ancient Greece and ancient India (Kosambi, 2005). However, their records have not survived; so the oldest available records now are those in Seoul, Korea, already presented in Section 2.02.1.3 and Figure 7 (upper), which go back to 1770, even though measurements must have been taken in much earlier periods since 1441 (Arakawa, 1956). The traditional device for rainfall measurement, known as rain gauge or pluviometer, is still in use today and, in fact, remains the most accurate device also providing the calibration basis for new measurement devices and techniques. It is a simple cylinder whose opening has an area (e.g., 200– 500 cm2 according to World Meteorological Organization (1983)) larger than (e.g., 10-fold) the cross section of the cylinder, which allows a greater sensitivity of the reading of the rainfall depth in a millimetric ruler attached to the cylinder. In another type of instrument, known as cumulative gauge, which is placed in inaccessible areas, the diameter of the cylinder may be larger than that of the opening, to enable the storage of a large volume of precipitation between the times of two visits to the place.

54

Precipitation

In an autographic (or recording) rain gauge, also known as a pluviograph, the water depth in the cylinder is recorded with the help of a mechanism involving a floating device. Another type of recording gauge, known as a tipping-bucket gauge, introduces the rainwater to one of a pair of vessels with a known small capacity (typically equivalent to 0.2 mm of rainfall) that is balanced on a fulcrum; when one vessel is filled, it tips and empties, while the time of this event is recorded, and the other vessel is brought into position for filling. In traditional autographic devices, these recordings are done on a paper tape attached to a revolving cylinder driven by a clockwork motor that is manually wound. In modern instruments, this device is often replaced by electronic systems, which provide digital recordings on a data logger and/or a computer connected by a cable or radio link. A rain gauge does not include all precipitation forms, snow in particular, except in light snowfalls when the temperature is not very low and the snow melts quickly. Generally, accurate measurement of snow precipitation (the water equivalent) needs specific instruments, equipped with a heating device to cause melting of snow. If such an instrument is not available, the snow precipitation is estimated as 1/10 of the snowfall depth (see justification in Section 2.02.1.3).

2.02.3.1.2 Typical processing of rain gauge data Measurement of precipitation in rain gauges is followed by several consistency checks to locate measurement errors and inconsistencies. Errors are caused due to numerous reasons, including human lapses and instrument faults, which may be systematic in case of inappropriate maintenance. Inconsistencies are caused by changes of installed instruments, changes in the environmental conditions (e.g., growing of a tree or building of a house near the rain gauge), or movement of the gauge to a new location. When errors are detected, corrections of the measurements are attempted. The standard meteorological practices include checks of outliers (a measured value is rejected if it is out of preset limits), internal consistency (checks are made whether different variables, e.g., precipitation and incoming solar radiation are compatible with each other), temporal consistency (the consistency of consecutive measurements is checked), and spatial consistency (the consistency of simultaneous measurements in neighboring stations is checked). Such checks are done in the timescale of measurement (e.g., daily for pluviometers or hourly for pluviographs) but systematic errors can only be located at aggregated (e.g., annual) timescales. The most popular method applied at an aggregated timescale for consistency check and correction of inconsistent precipitation data is that of the double-mass curve, which is illustrated in Figure 21. The method has a rather weak statistical background and is rather empirical and graphical (but there is a more statistically sound version in the method by Worsley (1983)). The double-mass curve is a plot of the successive cumulative annual precipitation Syi at the gauge that is checked versus the successive cumulative annual precipitation Sxi for the same period of a control gauge (or the average of several gauges in the same region). If the stations are close to each other and lie in a climatically homogeneous region, the annual values should correlate to each other. A fortiori, if the

two series are consistent with each other, the cumulated values Syi and Sxi are expected to follow a proportionality relationship. A departure from this proportionality can be interpreted as a systematic error or inconsistency, which should be corrected. Such a departure is usually reflected in a change in the slope of the trend of the plotted points. The aggregation of annual values xi and yi to calculate Sxi and Syi is typically done from the latest to the oldest year. Figure 21 (upper) shows the double-mass curve for 50 pairs of values representing annual precipitation at two points, whose cross-correlation (between xi and yi) is 0.82. The newest 25 points form a slope of m ¼ 0.70, whereas the oldest 25 form a much greater slope, m0 ¼ 0.95. Assuming that the newest points are the correct ones (with the optimistic outlook that things are better now than they were some years before), we can correct the older 25 annual yi by multiplying them with the ratio of slopes, l ¼ m/ m0 : ¼ 0.737. A second double-mass curve, constructed from the corrected measurements, that is, from y0i :¼ yi for ip25 (the newest years) and y0i :¼ lyi for i425 (the oldest years) is also shown in Figure 21 (upper). In fact, the data values used in Figure 21 are not real rainfall data but rather are generated from a stochastic model (Koutsoyiannis, 2000a, 2002) so that both stations have equal mean and standard deviation (1000 and 250 mm, respectively), be correlated to each other (with correlation coefficient 0.71) and, most importantly, exhibit HK behavior (with H ¼ 0.75, compatible with the values found in the real-world examples of Section 2.02.1.5). Hence, evidently, all values are correct, consistent, and homogeneous, because they were produced by the same model assuming no change in its parameters. Thus, the example illustrates that the method can be dangerous, as it can modify measurements, seemingly inconsistent, which however are correct. While this risk inheres even in time-independent series, it is largely magnified in the presence of HK behavior. Figure 21 (lower) provides a normal probability plot of the departure of the ratio l from unity (where the horizontal axis z is the standard normal distribution quantile and the distributions were calculated by the Monte Carlo method) for two cases: assuming independence in time and assuming HK behavior with H ¼ 0.75 as in the above example. The plots clearly show that, for the same probability, the departure of l from unity in the HK case is twice as high as in the classical independence case. For the HK case, departures of 70.25 from unity appear to be quite normal for 25-year trends and even more so for finer timescales, that is, 70.35 to 70.40 for 10-year to 5-year consecutive trends (not shown in Figure 21). Note that the method is typically applied even for corrections of as short as 5-year trends (Dingman, 1994), and so its application most probably results in distortion rather than correction of rainfall records. Apparently, the correction of the series using the doublemass curve method removes these trends that appear in one of the two time series. Removal of trends results in reduction of the estimated Hurst coefficient or even elimination of the exhibited HK behavior (Koutsoyiannis, 2003a, 2006b). Thus, if we hypothesize that the HK behavior is common in precipitation, application of methods such as the double-mass curves may have a net effect of distortion of correct data, based on a vicious circle logic: (1) we assume time independence of

Precipitation

55

50 000 Data Fitted broken line Adjusted data Adjusted straight line

40 000

Generation model Σyi (mm)

30 000

20 000

10 000

0 0

10 000

20 000 30 000 Σ xi (mm)

40 000

50 000

0.4 H = 0.5 (classical independence case) H = 0.75 (HK case)

Departure of slope ratio from 1, λ−1

0.3

0.2

0.1

0

−0.1

−0.2

−0.3 −3

−2

−1

0

1

2

3

Standard normal variate, z Figure 21 Illustration of the double-mass curve method and the associated risks in applying it. (Upper): Typical double-mass curve for 50 pairs of points, where the first 25 (newest) and the last 25 (oldest) form slopes m ¼ 0.7 and m0 ¼ 0.95, respectively; the adjusted points with l ¼ m/m0 ¼ 0.737 are also shown. (Lower): Comparison of probability distributions of the departure of the ratio l from unity for series independent in time or with HK behavior with H ¼ 0.75; the distributions were calculated using the Monte Carlo method based on synthetic series with a total size of 1000.

the rainfall process; (2) we interpret manifestation of dependence (the HK behavior in particular) as incorrectness of data; (3) we modify the data so as to remove the influence of dependence; and (d) we obtain a series that is much closer to our faulty assumption of independence. The widespread use of the double-mass curve method in routine processing of

precipitation time series may thus have caused enormous distortion of the real history of precipitation at numerous stations worldwide, in addition to masking HK behavior. The above discourse aims to issue a warning against unjustified use of consistency check and correction methods that could eliminate the extreme values (see, e.g., the note about

56

Precipitation

the Seoul station in Section 2.02.1.3) and the long-term variability implied by the HK behavior; the effect of both these mistreatments of data is a serious underestimation of design precipitation and flow in engineering constructions and management decisions. As a general advice for their correct application, we can stress that all methods of this type should never be applied blindly. An inspection of local conditions (environment of the rain gauge station and practices followed by the observer) as well as of the station’s archive history is necessary before any action is taken toward altering the data. Unless information on local conditions and archive history justify that inconsistencies or errors exist, corrections of data should be avoided.

2.02.3.1.3 Interpolation and integration of rainfall fields The interpolation problem, that is, the estimation of an unmeasured precipitation amount y from related precipitation quantities xi (i ¼ 1, y, n) is encountered very often in routine hydrologic tasks, such as the infilling of missing values of recorded precipitation at a station or the estimation of precipitation at an ungauged location. The integration problem refers to the estimation of an average quantity y over a specified area (or time period) based on measurements xi (i ¼ 1, y, n) of the same quantity and the same time period at different points (or respectively, at different time periods at the same point). The literature provides a huge diversity of methods, most of which, however, could be reduced to a linear statistical relationship applicable to both the interpolation and the integration problem:

y ¼ w1 x1 þ y þ wn xn þ e

ð28Þ

where wi denotes a numerical coefficient (weighting factor) and e denotes the estimation error. The same could be written in vector form:

y ¼ wTx þ e

ð29Þ

with w :¼ [w1, y, wn]T and x :¼ ½x1 ; ? ; xn T , and the superscript T denotes the transpose of a vector (or a matrix). The notation in Equations (28) and (29) suggests that x, y, and e are treated as random variables, even though this may not be necessary in some of the existing methods. All interpolation techniques provide a means for estimating the numerical coefficients wi, either conceptually or statistically, whereas the statistical methods provide, in addition, information about the error. Most commonly, the latter information includes the expected value me :¼ E½e and its standard deviation se :¼ ðVar½eÞ1=2 . A statistical estimation in which E½e ¼ 0 is called unbiased, and one in which the mean square error MSE :¼ E½e 2  ¼ s2e þ m2e , is the smallest possible, is called best; if both these happen, the estimation is called best linear unbiased estimation (BLUE). While the BLUE solution is in principle quite simple (see below), the estimation of its weighting factors is not always straightforward. Hence, several simplified statistical methods as well as empirical conceptual methods are in common use. Another reason that explains why such a diversity of methods has emerged is the different type of objects that each of the elements of the vector x represents. For instance, in temporal interpolation, these

elements can be observed values at times before and after the time of interpolation. In spatial interpolation these can be simultaneously observed values for stations lying in the neighborhood of the point of interpolation. Simultaneous temporal and spatial interpolation, although unusual, may be very useful. For example, an optimal way to infill a missing value in a time series at a specific time would be to include in x measurements taken in neighboring gauges at this specific time, as well as measurements taken at the point of interest at preceding and subsequent times. Let us first examine the different methods in which the estimation of y is based on a single observation x  xi at one neighboring (in space or time) point only. Here is a list of options, in which the following notation has been used: mx :¼ E½x and my :¼ E½y are the expected values of x and y, respectively; s2x :¼ E½ðx  mx Þ2  and a2y :¼ E½ðy  my Þ2  are the variances of x and y, respectively; sxy :¼ E½ðx  mx Þðy  my Þ is the covariance of x and y; and rxy :¼ sxy/(sx sy) is the correlation of x and y. 1. Equality: y ¼ x. The single point of observation considered in this naive type of interpolation is the station i nearest to the interpolation point, with x  xi. As discussed below, this simple interpolation forms the background of the Thiessen method of spatial integration. It is generally biased, with bias me ¼ my  mx and its MSE is s2y þ s2x  2sxy þ m2e . However, if the precipitation field is stationary (so that the means and variances at all points are equal to global parameters m and s2, respectively), it becomes unbiased, with MSE ¼ 2s2 (1  rxy). Evidently, for rxyo0.5, the method results in MSE 4 s2 and therefore there is no meaning in adopting it for low correlation coefficients (an estimate x ¼ m would be more effective). 2. Normal ratio: y ¼ w x with w ¼ my/mx. This is a better alternative to the equality case, but it requires a sample of measurements to be available for y in order to estimate the average my. This estimation is unbiased (me ¼ 0) but not best (MSE ¼ sy2 þ sx2 my2/mx2  2 sxy my/mx). 3. Homogenous linear regression: y ¼ wx with w ¼ E½y x= E½x 2  ¼ ðsxy þ mx my Þ=ðs2x þ m2x Þ. This is a biased estimation (me ¼ my  w mx) albeit best ðMSE ¼ s2y þ ðm2y s2x  2mx my sxy  s2xy Þ=ðm2x þ s2x ÞÞ. 4. Linear regression: y ¼ w x þ b with w ¼ Cov½y x=Var½x ¼ sxy =s2x and b ¼ my  w mx. This can be derived from Equation (28) by adding an auxiliary variable whose values are always 1 (i.e., y ¼ w x þ b 1). It has the properties of being both unbiased and best, with MSE ¼ s2y ð1  r2xy Þ. However, it has the deficiency of potentially resulting in negative values, if bo0, or of excluding values between 0 and b if b 4 0. Another drawback emerges when many values of y are estimated in an attempt to extend a record of y based on a longer record of x. In this case, the resulting extended record has negatively biased variance, because the method does not preserve variance. To remedy this, a random error e should be added (using the probability distribution of e), which however is not determined in a unique manner and makes the method no longer best. 5. Organic correlation: y ¼ wx þ b with w ¼ sign [rxy] sy/sx and b ¼ my  w mx. This preserves both mean (i.e., it is unbiased) and variance, but it is not best ðMSE ¼ 2s2y ð1  jrxy jÞ.

Precipitation Evidently, for |rxy|o0.5, the method results in MSE 4 s2y and therefore adopting it is pointless for low correlation coefficients. Similar to the standard linear regression, the organic correlation retains the deficiency of producing negative values or excluding some positive values. Coming to the interpolation based on multiple xi, in the simplest case, all weights wi are assumed equal for all i, that is, wi ¼ 1/n so that y is none other than the average of xi (the arithmetic mean method). This simple version is used very often to fill in sparse missing values of rain gauge records. The quantities xi could be simultaneous measurements at neighboring points (say, within a radius of 100 km), or at neighboring times, or both. Here, neighboring times should not necessarily be interpreted in the literal meaning, but with an emphasis on similarity of states. For example, a missing value of monthly precipitation in April 2000 could be estimated by, say, the average of the precipitation of the April months of 1998, 1999, 2001, and 2002. In another version, the average of all April months with available data are used, but a local average (as we have already discussed in Section 2.02.1.4) is preferable over an overall average, assuming that precipitation behaves like an HK process rather than a purely random one; this is similar to taking the average of points within a certain distance rather than a global average in spatial interpolation. This is not only intuitive but it can have a theoretical justification (D. Koutsoyiannis, personal notes), according to which for an HK process with H ¼ 0.7, a local average based on 3 time steps before and 3 after the interpolation time is optimal (produces lowest MSE); the optimal number of points becomes 2 þ 2 and 1 þ1 for H ¼ 0.75 and HX0.8, respectively. This simple method does not impose any requirement for calculation of statistical quantities for its application. Another method of this type, which takes account of the geographical locations and, in particular, the distances di between the interpolated stations, is the method of inverse distance weighting (IDW). In each of the basis stations, it assigns weights as

db wi ¼ Pk i b j¼1 dj

ð30Þ

where the constant b is typically assumed to be 2. Among methods whose application requires statistical quantities to be known, the simplest is a direct extension of the normal ratio method, in which wi ¼ ð1=nÞðmy =mxi Þ. The BLUE method itself belongs to this type. Initially, we can observe that a simple but biased solution for w in Equation (29) can be easily obtained as

w ¼ C 1 g;

me ¼ my  wT lx ;

s2e ¼ s2y  gT C1 g ¼ s2y  wT g

ð31Þ

where g :¼ Cov½y; x is the vector whose elements are the covariances of y with x (see Section 2.02.1.5) and C :¼ Cov½x; x is the positive definite symmetric matrix whose elements are the covariances of the vector x with itself. One way to make it unbiased is to add an auxiliary variable xnþ1 whose values are always 1. This is the multivariate extension of the typical linear regression described in point 4 of the previous list, and thus it retains the deficiency of potentially

57

producing negative values or excluding some positive values. A better way to make it unbiased is to add a constraint my ¼ wT lx (the bivariate analog of this is the equality case, described in point 1 of the list). In the latter case, the MSE becomes

MSE ¼ s2e ¼ s2y þ m2y þ wT ðC þ lx lTx Þw  2wT ðg þ my lx Þ

ð32Þ

Minimization of the MSE with the above constraint using a Lagrange multiplier –2l results in the system of equations

lTx w ¼ my

Cw þ lx l ¼ g;

ð33Þ

whose solution for the n þ 1 unknowns w1, y, wn, l is 0

w0 ¼ C 1 g0

ð34Þ

where

" 0

w :¼

# " w C 0 ; C :¼ l lTx

" # # lx g 0 ; g :¼ my 0

ð35Þ

The value of the error is then calculated as

MSE ¼ s2e ¼ s2y þ wT Cw  2wT g

ð36Þ

As seen in Equations (31) and (34), the application of the method requires a number of covariances to be estimated (specifically, this number is (n2 þ 3n)/2, given that C is symmetric). Not only does this restrict the method’s application to points where measurements exist, in order to estimate the covariances, but, when n is large, it is infeasible to reliably estimate so many parameters from data and to derive a positive definite C. The viable alternative is to assume a parametric stochastic model for the precipitation field. In the simplest case, the field could be assumed stationary and isotropic, where mxi ¼ my ¼ m; sxi ¼ sy ¼ s, and the covariance among any two points i, j is a function f of the geographical distance dij between these points, that is, sij : ¼ Cov[xi , xj ] ¼ f(dij). In this case, Equation (35) simplifies to

" 0

C :¼

C 1T

" # # " # 1 w g 0 0 ; w :¼ 0 ; g :¼ l 1 0

ð37Þ

where l0 ¼ l m and 1 is a vector with all its elements equal to 1. The last solution is widely known as ‘kriging’ (although kriging is sometimes formulated not in terms of covariance as in here, but in terms of the so-called semivariogram, a notion that is not appropriate for processes with HK behavior). We can observe from Equation (37) that the solution is now independent of m, as is also the error, which is still calculated from Equation (36). It only depends on the covariance function f(d). A function f(d) compatible with the HK behavior of precipitation, as discussed in Section 2.02.1.5, is of the form

f ðdÞ ¼ minðc; a d 4

H4

Þ

ð38Þ

where H is the Hurst coefficient and cb0 and a are parameters; in particular, c violates theoretical consistency but has

58

Precipitation

been introduced to avoid problems related to the infinite covariance for distance tending to zero. It can be observed that if the point of interpolation coincides with any one of the basis points i, then g is identical to one of the columns of C and g0 is identical to one of the columns of C0 . Thus, given the symmetry of C and C0 , from Equation (31) or (34), we obtain that w is a unit vector, that is, all elements are zero except one, which will be equal to 1. This shows the consistency of the method, that is, its property to reproduce the measurements at gauged points with zero error. All of the above methods that can interpolate at an arbitrary point (rather than only at a gauged one) provide a basis for numerical integration to find the average precipitation over a specific area A. Eventually, these methods result again in Equation (28) or (29), where now y is the areal average precipitation. In particular, in the arithmetic mean and the normal ratio methods, because they do not make any assumption about the position of the point to which interpolation refers, the estimate y is an interpolation at any point and a spatial average as well. The equality method works as follows: the geographical area of interest is divided into polygons, the so-called Thiessen polygons, each of which contains the points nearest to each of the stations. All points belonging to a specified polygon are regarded to have received a precipitation amount equal to that of the station corresponding to this polygon. Thus, in the integration, we use either Equation (28) or (29), where all gauged xi in the area are considered with weights wi ¼ Ai/A, whereas Ai and A are the areas of the polygon corresponding to xi. The remaining methods (IDW and BLUE) can be explicitly put in the form of Equation (29), but this is rather tedious if done analytically. A simpler alternative is to make interpolations to many points, for example, on a dense square grid. In turn, the gridded interpolations could be used for integration using equal weights for all grid points (i.e., arithmetic mean).

2.02.3.2 Radar Estimates of Precipitation Radio detection and ranging (radar) was developed at the beginning of World War II as a remote-sensing technique to measure the range and bearing of distant objects (such as ships and airplanes) by means of radio echoes (e.g., Battan, 1973). Since the early 1970s, radar techniques have also been used for the identification (i.e., shape, size, motion, and thermodynamic properties) of precipitation particles. The latter are weather-related distributed targets, which in contrast to ships and airplanes, have characteristics that evolve in time and depend on the atmospheric conditions. Because of their ability to provide estimates of areal precipitation quickly (i.e., at time intervals of about 5–15 min), at high resolutions (i.e., down to spatial scales of about 1 km) and over wide areas (i.e., with an effective range of about 200–400 km), radars have found wide application in atmospheric research, weather observation, and forecasting (e.g., Atlas et al., 1984; Doviak and Zrnic, 1993; Uijlenhoet, 1999, 2008; Bringi and Chandrasekar, 2001; Krajewski and Smith, 2002; Testik and Barros, 2007). An example is the next generation weather radar (NEXRAD) network with 159

operational weather surveillance radar 88 Doppler (WSR88D) units (as of February 2009), deployed throughout the continental United States and at selected locations overseas. According to NOAA’s weather service (US National Oceanic and Atmospheric Administration, 2009), since its establishment in 1988, the NEXRAD project has provided significant improvements in severe weather and flash flood warnings, air traffic control, and management of natural resources.

2.02.3.2.1 Basics of radar observation and measurement A typical weather radar has three main components (Battan, 1973): (1) the transmitter, which generates short pulses of energy in the microwave frequency portion of the electromagnetic spectrum, (2) the antenna, which focuses the transmitted energy into a narrow beam, and (3) the receiver, which receives the backscattered radiation from distant targets that intercept the transmitted pulses. Some important parameters, and their range of values, that characterize the radar equipment are (Rogers and Yau, 1996): (1) the instantaneous power of the pulse PtE10–103 kW (also referred to as peak power), (2) the duration of the pulse tE0.1–5 ms, (3) the frequency of the signal nE3–30 GHz, (4) the pulse repetition frequency (PRF) frE200–2000 Hz, defined as the reciprocal of the time interval tmax that separates two distinct pulses (i.e., tmax ¼ f 1 r E 0:5  5 ms), and (5) the beamwidth of the antenna yE11, defined as the angular separation between points where the power of the transmitted signal is reduced to half of its maximum value (or equivalently 3 dB below the maximum). The latter is attained at the beam axis. The wavelength l of the signal is defined as the distance between two sequential crests (or troughs) of the electromagnetic wave and it is related to its frequency as

ln ¼ c

ð39Þ

where c ¼ 3 108 m s1 is the velocity of light in vacuum. It follows from Equation (39) that typical frequencies n ¼ 3–30 GHz correspond to wavelengths l between 10 and 1 cm, but most weather radars operate at wavelengths l ¼ 3–10 cm (X-, C-, and S-band; see, e.g., Uijlenhoet and Berne, 2008). Shorter wavelengths are more effectively attenuated by atmospheric hydrometeors and precipitation particles (hence the transmitted signal has a small effective range), whereas for longer wavelengths the backscattered radiation from the precipitation particles does not have sufficient power to be detected by the receiver without noise induced by ground targets (e.g., Uijlenhoet, 2008). When conducting radar observations and measurements, the direction of the target is obtained from the azimuth and elevation of the antenna when the returning echo is received. The range r of the target is calculated from the relation

r ¼ c t=2

ð40Þ

where t is the time interval between the transmission of the pulse and the reception of the echo. If the target is moving, the radial velocity ur of the target (i.e., in the radar-pointing direction) can be obtained from the frequency shift Dn of the

Precipitation

received relative to the transmitted signal. The frequency shift is caused by the Doppler effect and it is related to ur as:

Dn ¼ 2ur =l

ð41Þ

with positive Dv being associated with targets that move toward the radar. If t (the time interval between transmission and reception) is larger than tmax (the reciprocal of the pulse repetition frequency, fr), the echo from the target will reach the receiver after a new pulse has been transmitted. Hence, targets that return enough energy to be detected by the receiver (see below) and are located at distances r4rmax ¼ c/(2fr), will appear unrealistically close to the antenna. Thus, rmax is the maximum range within which targets are indicated correctly on the radar screen and it is usually referred to as the unambiguous range (Battan, 1973; Rogers and Yau, 1996). The visibility of a target by the radar depends on whether the returning signal has sufficient power Pr to be detected by the receiver. As an example, we consider a point target (i.e., a target with linear dimension smaller than about 10% of l) with cross section At located at distance r from the radar. We suppose that the radar transmits pulses with peak power Pt that propagate isotropically in space (i.e., in a 3D sphere). It follows from simple geometric considerations that the power Pi intercepted by the target is

Pi ¼

Pt At 4pr 2

ð42Þ

where 4pr2 is the surface area of a sphere with radius r. If the transmitted signal is focused in a narrow beam by the antenna (as is commonly the case), Equation (42) becomes

Pi ¼ G

Pt At 4pr 2

ð43Þ

2

where G ¼ (4p Ae)/l is a dimensionless constant called the antenna axial gain that depends on the characteristics (i.e., the wavelength l) of the signal and the aperture Ae of the antenna. Assuming that the target scatters the intercepted signal isotropically in space, the power Pr that reaches the radar is

Pr ¼

Pi Ae Pt At Ae l 2G 2 ¼G 2 ¼ Pt At ð4pÞ 3 r 4 2 4pr 2 ð4pr Þ

ð44Þ

If the power Pr is large enough to be detected by the receiver without unwanted echoes (e.g., noise from ground targets), the target is visible to the radar and it is indicated on the radar screen. For nonisotropic scatterers, the cross section of the target At should be replaced by the backscatter cross section s of the target. For spherical particles with diameter Dol/10, usually referred to as Rayleigh scatterers, s can be calculated from the relation (Battan, 1973)

s¼

p 5 jKj 2 D 6 l4

ð45Þ

59

where |K| is the amplitude of the complex refraction index (|K|2E0.93 for liquid water and 0.21 for ice), which characterizes the absorptive and refractive properties of the spherical scatterer. Due to the much higher value of |K|2 for liquid water relative to ice (about 4.5 times higher), the melting layer of ice particles in precipitation-generating weather systems appears on the radar screen as a bright band of high reflectivity.

2.02.3.2.2 Radar observation of distributed targets and the estimation of precipitation For a typical weather radar that operates in the C-band portion of the electromagnetic spectrum (l ¼ 3.75–7.5 cm), raindrops and snowflakes (i.e., particles with effective diameters Do5– 6 mm) can be approximated as Rayleigh spherical scatterers with backscatter cross section s given by Equation (45). However, there are reasons why atmospheric hydrometeors should not be treated as isolated point targets. One reason is that the pulse transmitted by the radar illuminates simultaneously, numerous precipitation particles that are included in a certain volume of air V, referred to as the resolution volume of the radar. Hence, the returned signal contains spatially averaged information from the whole population of raindrops and snowflakes in V. For parabolic antennas, where the beam pattern is approximately the same in all directions, an accurate estimate of V can be obtained by assuming that the resolution volume is a cylinder with effective height equal to half of the pulse length l ¼ c t and diameter dV ¼ r y, that is, the separation distance between points where the power of the transmitted signal is reduced to half of its maximum value. This gives

 2 ry ct V¼p 2 2

ð46Þ

where y is in radians. Equation (46) assumes that all energy in the radar transmitted pulse is contained within the half-power beamwidth; assuming a Gaussian shape of the beam pattern, the denominator of (46) (and, likewise, that of (49) below) should be multiplied by a factor 2 ln 2 (Probert-Jones, 1962). Another reason why raindrops and snowflakes cannot be treated as isolated point targets is that their turbulent motion that causes the power Pr of the returned signal to fluctuate in time. To this extent, an accurate approximation of the timeaveraged power Pr (over a sufficiently long interval of about 102 s), which accounts also for multiple backscattering cross sections, is given by (Rogers and Yau, 1996)

l 2G 2 X s Pr ¼ Pt ð4pÞ 3 r 4 V

ð47Þ

where r is the time-averaged range of the resolution volume V, and the summation is taken over all s in V. For Rayleigh scatterers, Equations (45) and (47) are combined to give

p 2 G 2 jKj 2 X 6 Pr ¼ Pt D 64r 4 l 2 V

ð48Þ

60

Precipitation

Assuming homogeneity of the population of hydrometeors in V, Equation (48) can be written as

Z N p 2 G 2 jKj 2 Pr ¼ Pt V nðDÞD6 dD 64r 4 l 2 0 p 3 G 2 jKj 2 y 2 ct jKj 2 Z ¼ Pt Z¼C 2 2 2 512r l r

ð49Þ

where n(D) is the size distribution of precipitation particles in V (i.e., number of particles per unit diameter and per unit volume of air), C is the so-called radar constant that depends solely on the characteristics of the system under consideration, and

ZN Z :¼

nðDÞD 6 dD

ð50Þ

0

is the reflectivity factor with units (length3) that depend solely on the size distribution of the precipitation particles. For the Marshall and Palmer (1948) parametrization described by Equation (24), Equation (50) takes the form

Z ¼ 720 n0 b7

ð51Þ

where n0 and b are the intercept and scale parameters of the exponential size distribution. For the expressions given in Section 2.02.2.3.2, for rain and snow we obtain

ðaÞ Z ¼ 296 i 1:47 ðrainÞ ðbÞ Z ¼ 3902 i 2:49 ðsnowÞ

ð52Þ

where Z has units of mm6 m3 and i is the rainfall intensity (or the water equivalent of the accumulated snow at ground level) in millimeters per hour. For rain, Equation (52a) is very close to the empirical Z  i relationships (usually referred to as Z  R relationships, where Ri denotes the rainfall intensity) found in the literature (e.g., Marshall et al., 1955; Battan, 1973; Uijlenhoet, 1999, 2001, 2008), whereas for snow there is more variability and Equation (52b) should be seen only as an approximation. When combined, Equations (40), (49), and (52) allow conversion of radar measurements (i.e., Pr , t and r) to precipitation intensity i.

2.02.3.3 Spaceborne Estimates of Precipitation The history of observation of Earth from space started on 4 October 1957, when the Soviet Union successfully launched Sputnik-I, the first artificial satellite. Sputnik-I provided information on the density of the highest layers of the atmosphere and on the radio-signal distribution in the ionosphere. The first launch was immediately followed by the launch of Sputnik-II by the Soviet Union on 3 November 1957 and the launches of Explorer-I (1 February 1958), Vanguard-I (17 March 1958), Vanguard-II (17 February 1959), and TIROS-I (1 April 1960) by the United States of America. The success of TIROS-I in surveying atmospheric conditions (in particular, the cloud coverage of Earth) opened a new era for

meteorological research and development using spaceborne observations. Since the 1970s, meteorological satellites have become essential in studying the development and evolution of weatherrelated phenomena over the 71% of the Earth’s surface covered by sea, where other types of measurements are unavailable. For example, the Tropical Rainfall Measuring Mission (TRMM; Simpson et al., 1988; Kummerow et al., 1998), which started in November 1997 by the National Aeronautics and Space Administration (NASA) of the United States and the National Space Development Agency (NASDA) of Japan, has provided vast amounts of rainfall and energy estimates in tropical and subtropical regions and advanced the understanding and modeling of extreme rainfall events caused by tropical cyclones (e.g., Lonfat et al., 2004, 2007; Chen et al., 2006, 2007; Langousis and Veneziano, 2009a, 2009b). TRMM data have also been used to improve the accuracy of high-resolution weather forecasts produced by limited-area models (e.g., Lagouvardos and Kotroni, 2005) and to investigate the relationship between lighting activity, microwave brightness temperatures (see below), and spaceborne radar-reflectivity profiles (Katsanos et al., 2007). We can distinguish two types of sensing by satellites, passive and active. Passive sensing is based on measuring the radiative intensity emitted or reflected by particles in the atmosphere, such as cloud droplets and hydrometeors of precipitable size. Active sensing is conducted using radar equipment carried by the satellite. Next, we discuss some basic principles of passive remote sensing in the visible (V, lE0.39– 0.77 mm), IR (wavelengths lE0.77 mm–0.1 mm), and microwave (MW, lE0.1 mm–10 cm) portions (channels) of the electromagnetic spectrum. The basic principles of operation of active sensors are similar to those of radars, reviewed in Section 2.02.3.2. For a more detailed review on the principles and techniques of remote sensing, the reader is referred to Barrett and Martin (1981), Elachi (1987), Stephens (1994), and Kidder and Vonder Haar (1995).

2.02.3.3.1 The IR signature of cloud tops The high absorptivity of cloud droplets in the IR spectral range causes clouds to appear opaque in the IR channel. Hence, the IR radiation received by the satellite’s radiometer originates mostly from the cloud tops, which can be approximated with sufficient accuracy as black bodies, that is, as objects that absorb all incident radiation and emit it at a rate that depends solely on their temperature. In this case, we can use Stefan– Boltzman’s law of radiation (e.g., Barrett and Martin, 1981) to calculate the temperature Tb of the cloud tops from the intensity J of the received IR radiation:

Tb ¼ ðJ=sSB Þ1=4

ð53Þ

where sSB ¼ 5.7 108 W m2 K4 is the Stefan–Boltzman constant and Tb is in kelvins. Tb is usually referred to as brightness temperature (e.g., Smith, 1993) and, for a given atmospheric lapse rate g (see Section 2.02.2.1), it can be used to calculate cloud top heights. Evidently, lower brightness temperatures Tb correspond to clouds with higher tops and larger probabilities of rain.

Precipitation

Hence, we can develop regression equations to relate brightness temperatures to observed surface rainfall rates (e.g., Griffith et al., 1978; Stout et al., 1979; Arkin, 1979; Richards and Arkin, 1981; Arkin and Meisner, 1987; Adler and Negri, 1988). Two important limitations apply (Richards and Arkin, 1981; Liu, 2003): (1) due to the statistical character of the regressed quantities, the accuracy of the rainfall-retrieval algorithm increases with increasing scale of spatial or temporal averaging, and (2) the parameters of the regression depend on the climatology of the region and, therefore, cannot be used at regions with different climatic characteristics. An example of surface rainfall estimation from IR images is the temperature threshold method developed by Arkin (1979), Richards and Arkin (1981) and Arkin and Meisner (1987). Arkin (1979) used IR imagery from the Synchronous Meteorological Satellite-1 (SMS-1) and radar data from Global Atmosphere Research Program (GARP) Atlantic Tropical Experiment (GATE) to investigate the correlation between radarestimated precipitation rates and the fraction of areas with brightness temperature Tb below a certain threshold Tmin. The study found a maximum correlation (around 0.85) for a brightness temperature threshold TminE235 K (  38 1C). Richards and Arkin (1981) showed that a linear relationship is sufficient to describe the dependence between spatially averaged surface rainfall and the fraction of areas with Tbo235 K, with error variance that increases with decreasing scale of spatial averaging. Based on these results, Arkin and Meisner (1987) suggested the use of the Geostationary Operational Environmental Satellite (GOES) Precipitation Index (GPI) to calculate spatial rainfall averages in the tropics:

GPI ¼ 3ðmm=hÞFc H

ð54Þ

where GPI is the spatially averaged rainfall accumulation in a grid box of 2.51 latitude  2.51 longitude, Fc is the mean fraction (a dimensionless quantity between 0 and 1) of the grid box covered by brightness temperatures Tbo235 K, and H is the length of the observation period in hours. The temperature threshold method of Arkin (1979), Richards and Arkin (1981), and Arkin and Meisner (1987) produces accurate estimates of the spatially averaged rainfall in the tropical belt (301 S to 301 N), at grid scales larger than 2.51(E275 km) (Arkin and Meisner, 1987) and for averaging durations greater than about a month (Ba and Nicholson, 1998). The error increases significantly as we move to midlatitudes, especially during cold seasons (e.g., Liu, 2003). Extensions of the method include the use of the upper tropospheric humidity (UTH) in the vicinity of convective clouds as an additional predictive variable (Turpeinen et al., 1987), and the combination of IR and visible imagery (i.e., bi-spectral methods; see below) to exclude nonprecipitating clouds with high tops.

2.02.3.3.2 The visible reflectivity of clouds The signature of Earth in the visible (V) channel is due to the reflection of the sunlight by clouds and, when the sky is clear, the surface features. Consequently, visible imagery is available only during daylight hours. Due to its shorter wavelength, visible radiation can penetrate deeper into clouds than the

61

infrared portion of the electromagnetic spectrum, but similar to the IR channel, it still represents the upper portion of clouds and serves as an indirect signature of surface rainfall. However, visible reflectivity can complement the IR brightness temperatures to allow better classification of clouds and qualitative assessment of the probability of precipitation. This is the basis of the well-known bi-spectral methods (e.g., Lovejoy and Austin, 1979; Bellon et al., 1980; Tsonis and Isaac, 1985; Tsonis, 1987; O’Sullivan et al., 1990; Cheng et al., 1993; Cheng and Brown, 1995; King et al., 1995; Liu, 2003). The visible reflectivity of clouds increases fast with the increasing liquid water path, that is, the vertically integrated liquid water in the atmospheric column. Hence, we can use IR brightness temperatures to calculate the altitude of the cloud tops and visible reflectivities to obtain a qualitative estimate of the vertically averaged liquid water of the cloud, which is indicative of the rainfall potential. For example, low brightness temperatures (i.e., cold cloud tops) and high visible reflectivities (i.e., thick clouds) indicate cumulonimbus formations with high probability of precipitation (see Section 2.02.2.4.1), warm cloud tops and high visible reflectivities indicate stratiform rainfall (see Section 2.02.2.4.2), whereas cold cloud tops and low visible reflectivies indicate cirrus clouds, which are usually nonprecipitating. An example of bi-spectral methods is the RAINSAT technique developed by Lovejoy and Austin (1979) and Bellon et al. (1980). This technique uses visible reflectivities to reduce the number of false alarms obtained from the IR channel and more accurately estimate surface rainfall rates. The RAINSAT method was developed using GOES infrared and visible imagery and radar data from tropical (i.e., GATE) and mid-latitude (i.e., McGill weather radar, Quebec, Canada) locations as ground truth. The method was optimized by Cheng et al. (1993) and Cheng and Brown (1995) for the area of the UK, using IR and visible imagery from the European geostationary satellites Meteosat-2, Meteosat-3, and Meteosat-4 and rainfall retrievals from nine weather radars located in the United Kingdom and Ireland. A similar cloud classification technique has been proposed by Tsonis and Isaac (1985) and Tsonis (1987). This technique is based on cluster analysis of pixels with different brightness temperatures and visible reflectivities and has been developed using GOES satellite data and rainfall retrievals from the Woodbridge weather radar in Ontario, Canada.

2.02.3.3.3 The microwave signature of precipitation Contrary to the IR and visible spectral ranges, microwave radiation can effectively penetrate through cloud and rain layers and provide the signature of the integrated contribution of precipitation particles in the atmospheric column. Hence, brightness temperatures obtained from the MW channel are better linked to surface rainfall rates than the visible reflectivities and IR brightness temperatures. The type and size of the precipitation particles detected by the microwave radiometer depends on the frequency of the upwelling radiation. Above 80 GHz (i.e., wavelengths lo3.75 mm), ice crystals scatter the upwelling MW radiation and fade the signature of raindrops. Hence, above 80 GHz, the radiometer senses only ice, where lower brightness

62

Precipitation

temperatures are associated with more scattering, larger ice particles, and higher precipitation intensities at ground level. Below about 20 GHz (i.e., l41.5 cm), the radiative intensity of raindrops dominates the microwave signature of hydrometeors in the atmospheric column, whereas ice particles are virtually transparent. Thus, below 20 GHz, the microwave radiometer detects the vertically integrated signature of rainwater, where higher brightness temperatures are associated with more intense rainfall at ground level. Lowfrequency microwave imagery is especially useful when calculating surface rainfall rates over oceans, where the almost constant sea surface temperature and emissivity allow translation of the spatial and temporal variations of brightness temperatures to variations of sea-level rainfall rates (e.g., Liu, 2003). The same is not true over land, where the surface features cause the ground temperature and emissivity to vary significantly in space and time. Another limitation of lowfrequency microwave images is the saturation of the microwave channel at high rainfall rates, which causes negative biases of the obtained rainfall intensity (e.g., Liu, 2003; Viltard et al., 2006). Between 20 GHz and 80 GHz, scattering and emission by raindrops and ice particles occur simultaneously and the microwave radiation undergoes multiple transformations. Hence, the microwave radiometer detects different rain paths at different microwave frequency ranges. Combining brightness temperatures from different MW channels to more accurately assess surface rainfall rates is an open research problem and it has driven the development of many rainfall-estimation algorithms (Grody, 1991; Spencer et al., 1989; Alishouse et al., 1990; Berg and Chase, 1992; Hinton et al., 1992; Liu and Curry, 1992, 1993; Ferriday and Avery, 1994; Petty, 1994a, 1994b, 2001a, 2001b; Kummerow and Giglio, 1994a, 1994b; Ferraro and Marks, 1995; Kummerow et al., 1996, 2001; Berg et al., 1998; Aonashi and Liu, 2000; Levizzani et al., 2002). For a review of microwave methods of estimation over ocean and land, and their advantages and limitations, the reader is referred to Wilheit et al. (1994), Petty (1995), and Kidd et al. (1998) respectively.

2.02.4 Precipitation modeling As already clarified in Section 2.02.1.5, modeling of precipitation is not possible without using any type of a stochastic approach. Even the deterministic numerical weather forecast models, which determine the state and motion in the atmosphere by solving differential equations, to model precipitation, use parametrization schemes. These schemes, instead of describing the detailed dynamics of the precipitation process, establish and use equations of statistical type to quantify the output of the dynamical system. In addition, as mentioned in Section 2.02.1.5, the modern framework for predicting precipitation particularly as input to hydrological models (the ensemble forecasting), is of the Monte Carlo or stochastic type. The description of these stochastic techniques belongs to the sphere of weather forecasting and is not within the scope of this chapter. In more engineering-oriented applications, precipitation is typically modeled as an autonomous process, without particular reference to the atmospheric

dynamics. Next, we outline some of the most widespread modeling practices for precipitation but without details and mathematical formulations, which the interested reader can find in the listed references.

2.02.4.1 Rainfall Occurrence From the early stages of the analysis of precipitation intermittency, it was recognized that rainfall occurrences are not purely random. In other words, rainfall occurrence cannot be modeled (effectively) as a Bernoulli process in discrete time or, equivalently, as a Poisson process in continuous time. It should be recalled that in a Bernoulli process, an event (rainfall/wet state) occurs with a probability p (and does not occur with probability 1  p) constant in time, and each event is independent of all preceding and subsequent events. In a Poisson process, the times of occurrence of events (i.e., the starting times of rainfalls) are random points in time. In this process, the time differences between consecutive occurrences are independent identically distributed (IID) with exponential distribution. Both discrete time and continuous time representations of the rainfall occurrence process, which in fact are closely related (e.g., Foufoula-Georgiou and Lettenmaier, 1986; Small and Morgan, 1986), have been investigated. The most typical tool of the category of discrete time representations is the Markov chain model (Gabriel and Neumann, 1962; Feyerherm and Bark, 1964; Hershfield, 1970; Todorovic and Woolhiser, 1975; Haan et al., 1976; Chin, 1977; Katz, 1977a, 1977b; Kottegoda and Horder, 1980; Roldan and Woolhiser, 1982). In this model, any time interval (e.g., day) can be in one of two states, dry or wet, and it is assumed that the state in a time interval depends on the state in the previous interval. It was observed, however, that Markov chain models yield unsatisfactory results for rainfall occurrences, especially for dry intervals (De Bruin, 1980). Moreover, the interannual variance of monthly (or seasonal) total precipitation is greater than that predicted by Markov chain models, an effect usually referred to as overdispersion (Katz and Parlange, 1998). Extended versions of the binary state Markov chains using a higher number of past states may improve performance. Additional states in such model versions have been defined based on a combination of states of two consecutive periods (Hutchinson, 1990) or on accounting for the rainfall depth of each interval (Haan et al., 1976). A more effective enhancement is to use transition probabilities taking into account more than one previous interval, which leads to stochastic binary chains of order higher than one (Pegram, 1980; Katz and Parlange, 1998; Clarke, 1998). In more recent developments, to account for a long number of previous time intervals and simultaneously avoid an extremely high number of transition probabilities, it was proposed that, instead of the sequence of individual states of these intervals, one could use conditional probabilities based on aggregation of states of previous intervals (Sharma and O’Neill, 2002). Similarly, one could use a discrete wetness index based on the number of previous wet intervals (Harrold et al., 2003). An extension of the Markov chain approach to multiple sites has been studied by Pegram and Seed (1998).

Precipitation

In a more recent study, Koutsoyiannis (2006a) used the principle of maximum entropy, interpreted as maximum uncertainty, to explain the observed dependence properties of the rainfall-occurrence process, including the overdispersion or clustering behavior and persistence. He quantified intermittency by the probability p(1) that a time interval of length 1 h is dry, and dependence by the probability that two consecutive intervals are dry, that is by p(2), where in general p(k) denotes the probability that an interval of length k is dry. Using these two probabilities and a multiscale entropy-maximization framework, he was able to determine any conditional or unconditional probability of any sequence of dry and wet intervals at any timescale. Thus, he described the rainfall occurrence process including its dependence structure at all scales using only two parameters. The dependence structure appeared to be non-Markovian, yet not over-exponential. Application of this theoretical framework to the rainfall data set of Athens indicated good agreement of theoretical predictions and empirical data at the entire range of scales for which probabilities dry and wet can be estimated (from 1 h to several months). An illustration is given in Figure 22. In the continuous time representation of the rainfall occurrence process, the dominant tools are the cluster-based point processes (Waymire and Gupta, 1981a, 1981b, 1981c). These are essentially based on the prototype of the spatial distribution of galaxies devised by Neyman and Scott (1952) to describe their property of clustering relative to the Poisson process. With reference to storms, if they were regarded as instantaneous pulses positioned at random points in time, the logarithm of probability that the interarrival time exceeds a value x, or the log survival function, would be proportional to

63

x. However, empirical evidence suggests that the log survival function is a nonlinear concave function of x, which indicates a tendency for clustering of rainfall events relative to the Poisson model (Foufoula-Georgiou and Lettenmaier, 1986). This clustering has been modeled by a cascade of two Poisson processes, corresponding to two characteristic timescales of arrivals of storms and storm cells. The Neyman–Scott process with instantaneous pulses was the first one applied to rainfall occurrence (Kavvas and Delleur, 1981; Rodriguez-Iturbe et al., 1984), later succeeded by the Neyman–Scott rectangular pulses and the very similar Bartlett–Lewis rectangular pulse models (Rodriguez-Iturbe et al., 1987). The Bartlett–Lewis rectangular pulse model, which is the most typical and successfully applied model of this type, assumes that rainfall occurs in the form of storms of certain durations and that each storm is a cluster of random cells. The general assumptions of the rainfall occurrence process are: 1. Storm origins ti occur according to a Poisson process with rate l. 2. Origins tij of cells of each storm i arrive according to a Poisson process with rate b. 3. Arrivals of each storm i terminate after a time vi exponentially distributed with parameter g. 4. Each cell has a duration wij exponentially distributed with parameter Z. In the original version of the model, all model parameters are assumed constant. In a modified version, the parameter Z is randomly varied from storm to storm with a gamma distribution with shape parameter a and scale parameter n.

p(k)

1

0.1

0.01 1

10

100 k

1000

10 000

Figure 22 Probability dry p(k) vs. scale k (in h), as estimated from a hourly rainfall data set in Athens, Greece, and predicted by the maximum entropy model in Koutsoyiannis (2006a) for the entire year (circles and red full line) and the dry season (June–September; diamonds and blue full line). The model was fitted using two data points in each case (marked in full in the plot), that is, the probability dry for 1 h, pp(1), and 2 h, p(2), which are respectively 0.9440 and 0.9335 for the entire year and 0.9888 and 0.9860 for the dry season. The final model is expressed as 1=Z Z p ðk Þ ¼ p ½1þðx 1Þðk 1Þ , where the parameters are respectively Z ¼ 0.63 and x ¼ 0.816 for the entire year and Z ¼ 0.83 and x ¼ 0.801 for the dry season. For comparison, lines resulting from the Markov chain model are also plotted (dashed lines). From Koutsoyiannis D (2006a) An entropicstochastic representation of rainfall intermittency: The origin of clustering and persistence. Water Resources Research 42(1): W01401 (doi:10.1029/ 2005WR004175).

64

Precipitation Storms

Intensity (mm h−1)

4

Storm 1 Storm 2 Storm 3 Storm 4

3

2

1

0 2

1 Time (days)

Figure 23 Simulated realization of a series of four storms from the Bartlett–Lewis rectangular pulse model (modified version with randomly varying Z) occurring within three days (notice the overlap of storms 1 and 2, which is allowed by the model), implemented by the Hyetos software (see Section 2.02.4.4). The model parameters are l ¼ 0.94 d1, k ¼ b/Z ¼ 1.06, j ¼ g/Z ¼ 0.059, a ¼ 2.70, n ¼ 0.0068 d1, and mx ¼ sx ¼ 24.3 mm d1.

Subsequently, parameters b and g also vary so that the ratios k: ¼ b/Z and j: ¼ g/Z are constant. A major problem of these models was their inability to reproduce the probability of zero rainfall at multiple timescales (Velghe et al., 1994). In this respect, Foufoula-Georgiou and Guttorp (1986) noted that the Neyman–Scott model parameters are scale dependent and thus cannot be attributed a physical meaning. To ameliorate this, modifications of both the Neyman–Scott model (Entekhabi et al., 1989) and the Bartlett–Lewis model (Rodriguez-Iturbe et al., 1988; Onof and Wheater, 1993, 1994) were proposed. These are in fact based on the randomization of the mean interarrival time of one of the two Poisson processes. Evaluation and comparison of several cluster-based rectangular pulse models for rainfall were done by Velghe et al. (1994) and Verhoest et al. (1997), whereas a comprehensive review of Poisson-cluster models has been provided by Onof et al. (2000). An extension of the concept introducing a third Poisson process was proposed by Cowpertwait et al. (2007).

2.02.4.2 Rainfall Quantity In the discrete time representations of rainfall occurrence, the rainfall quantity in each wet interval is modeled separately from the occurrence process, usually based on statistical analysis of the observed record. In the point-process representations, the storms and cells are abstract quantities that do not fully correspond to real-world objects. Therefore, they cannot be identified in the recorded time series. An assumption is typically made that each cell has a uniform intensity xij with a specified distribution, and based on all assumptions, the statistical characteristics of the rainfall process at one or more timescales are derived analytically (Rodriguez-Iturbe et al., 1987, 1988). These statistical characteristics are compared to the empirically derived statistics, and, by minimizing the departures of the two, the model parameters are estimated. The distribution of the uniform intensity xij is typically assumed to be exponential with parameter 1/mx. Alternatively, one can choose a two-parameter gamma distribution with mean mx

and standard deviation sx. In this manner, the point-process models describe the entire rainfall process, including occurrence and quantity. A demonstration of the model is shown in Figure 23. However, in some cases (e.g., Gyasi-Agyei and Willgoose, 1997), point processes have been used to simulate merely rainfall occurrences and then have been combined with other models that simulate rainfall depths. Other modeling approaches for the rainfall process (including its intermittency) are reviewed in Srikanthan and McMahon (2001). With their typical assumptions, including those of the exponential or gamma distribution for rain-cell amount, the point-process models, despite providing satisfactory representation of the process at a specific timescale or a small range of timescales, cannot really perform satisfactorily over a wide range of scales and also lead to exponential distribution tails, whereas it has been recently recognized that the tails must be of power type (see Sections 2.02.1.5 and 2.02.5.2). Generally, the distribution function of rainfall varies among different timescales. At very fine scales, the density is J-shaped, that is, with a mode at zero, and perhaps with density tending to infinity as the rainfall depth or the intensity tends to zero. At coarse timescales such as monthly (for wet months) and annual, the distribution becomes bell-shaped and tends to become normal as the scale increases. However, its tail always departs from the exponential tail of the normal distribution. In fact, for theoretical reasons, if at the right tail, the survival function is a power function of the rainfall depth or intensity x, with exponent 1/k, that is, F*(x)px1/k (see Equation (18)), then it will be of the same type and will have precisely the same exponent 1/k at any timescale (the proof is omitted). This behavior of the tail is perhaps the only invariant distributional property across all scales, whereas the shape of the body of the distribution varies significantly across different scales. However, even this variation must have a simple and unique explanation, which is the principle of maximum entropy. Specifically, Koutsoyiannis (2005a) has shown that all diverse shapes of the distribution across different scales can be derived from the principle of maximum entropy constrained on known mean and variance.

Precipitation

Papalexiou and Koutsoyiannis (2008) proposed a single distribution (a power-transformed beta prime distribution, also known as generalized beta of second kind; see also Koutsoyiannis, 2005a) with four parameters, which provides good fits for rainfall intensity at timescales from hourly to annual. Only one of the four parameters (corresponding to the exponent of the tail) is invariant across scales. If the range of scales of interest is smaller, then specific special cases of this distribution can be used as good approximations. For example, the three-parameter Burr type VII distribution, which has the advantage of providing a closed form of the quantile function, can be used effectively for timescales from a few minutes to a couple of months (Papalexiou and Koutsoyiannis, 2009).

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Rainfall cells

2.02.4.3 Space–Time Models Space–time modeling of precipitation is one of the most demanding tasks of stochastic modeling in hydrology and geophysics. Rainfall intermittency should be modeled in both space and time, along with the motion of rainfall fields, the rainfall quantity, and its temporal and spatial structure. One of the relatively simple solutions has been provided by the extension of point-process models used for the rainfall process at a single site. This extension introduces a description of rainfall cells in space, in addition to that in time, and a motion of the cells. As an example, we summarize here the Gaussian displacement spatial–temporal rainfall model (GDSTM; Northrop, 1996, 1998). This model, is a spatial analog of a point-process model having a temporal structure similar to that of the Bartlett-Lewis rectangular pulse model described above and a spatial structure known as the Gaussian displacement structure, introduced by Cox and Isham (1988). Similar to its single-site analog, GDSTM assumes that rainfall is realized as a sequence of storms, each consisting of a number of cells. Both storms and cells are characterized by their centers, durations, and areal extents (see sketch in Figure 24) and, in addition, cells have certain uniform rainfall intensity. Specifically, the following assumptions characterize storms and cells. Storm centers arrive according to a homogeneous Poisson process of rate l in 2D space (denoted by x, y) and time (denoted by t) and move with a uniform velocity (Vx, Vy). Each storm has a finite duration L (assumed exponentially distributed with parameter b ¼ 1/mL) and an infinite areal extent, represented by an elliptical geometry with eccentricity E and orientation y, and incorporates a certain number of rainfall cells. However, a storm can be assigned a finite storm area, the area that contains a certain percentage of rainfall cells. The storm area varies randomly and in each storm, it is determined in terms of the realization of a random variable w, which determines uniquely (for the specific storm) a set of parameters s2x ; s2y , and r that determine the displacement of cell centers from the storm center. Specifically, w is Gammadistributed with shape and scale parameters determined in terms of the eccentricity e and the mean storm area ms. At the same time, the parameter r is determined in terms of the eccentricity e and the storm orientation, y. Following the generation of w, the parameters s2x and s2y are determined

Storm center Figure 24 Sketch of the spatial structure of the Gaussian displacement spatial–temporal rainfall model.

in terms of the eccentricity e, the storm orientation y, and the value of w. Each rainfall cell is assigned a center ðxc ; yc ; tc Þ. The time origin tc follows a Poisson process starting at the time ordinate of the storm origin t0 (with the first cell being located at this point) and ending at t0 þ L. The expected number of cells within that time interval is mc ¼ 1 þ b/g, where g is the cellgeneration Poisson process parameter. The spatial displacements from the storm center are random variables jointly normally distributed with zero means, variances s2x and s2y , and correlation r. Given these parameters, the displacement Dx of each cell is generated as a normal variate (0, sx) and the displacement Dy as a normal variate (my|x, sy|x). Furthermore, each cell has a finite duration D (assumed exponentially distributed with parameter 1/mD) and an elliptical area with major axis a, forming an angle y with the x-axis (west–east), pffiffiffiffiffiffiffiffiffiffiffiffiffi ffi and minor axis b ¼ 1  e 2 a. It is assumed that a is a random variable gamma distributed with shape and scale parameters depending on the mean storm area mA and the eccentricity e, respectively. Finally, each cell has an intensity x independent of any other variable, exponentially distributed with parameter 1/mx. The model is defined in terms of 11 independent parameters, namely: (1) the rate of storm arrival (number of storms per area per time), l; (2) the mean cell duration, mD; (3) the mean storm duration, mL; (4) the mean cell area, mA; (5) the mean storm area, ms; (6) the mean number of cells per storm, mc; (7) the mean cell intensity, mx; (8 and 9) the components of the cell and storm velocity in the x direction (east), Vx, and in the y direction (north), Vy; (10) the cell and storm eccentricity, e; and (11) the cell and storm orientation, y. Similar to its single-point analog, the entities of the spatial point-process model are abstract. To make the model outputs comparable to reality, integration from continuous time over a specific timescale and/or spatial scale is needed, from which the first- and second-order rainfall statistics are

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calculated. The latter serve as the basis for parameter estimation using either rain gauge or radar data. Due to model complexity, the calculation of the statistics can be done only numerically; hence, the entire model application (and the parameter estimation in particular, which needs numerical optimization, e.g., using the generalized reduced-gradient method) is laborious.

2.02.4.4 Rainfall Disaggregation and Downscaling Both disaggregation and downscaling refer to the generation of a precipitation field at a specific temporal and/or spatial scale given a known precipitation field (measured or simulated) at a certain larger temporal and/or spatial scale (lower resolution). Disaggregation and downscaling are very useful procedures and have several applications, such as in the following cases: 1. Global-scale weather-prediction models provide rainfall forecasts at a low resolution, for example, grid size of 50 km. Hydrologic models require the description of the precipitation field at a much higher resolution, with grid size of the order of 1 km. 2. Satellite precipitation estimates are available at a spatial scale X0.251 (latitude and longitude), or about 28 km at the equator, and a temporal scale of 3 h. Again, hydrologic applications require higher resolutions. 3. The majority of historical point rainfall records come from daily rain gauges, which have often been operational for several decades. The number of rain gauges providing hourly or sub-hourly resolution data is smaller by about an order of magnitude. However, hydrologic applications, especially flood studies, usually need hourly or even subhourly data. 4. In complex problems of stochastic generation of precipitation time series or precipitation fields, it is difficult to reproduce simultaneously, the long-term and the shortterm stochastic structure of precipitation using a single model. A better approach is to couple several models, starting from a large-scale model to represent the long-term behavior. The outputs of the latter are then disaggregated into finer scales. Note, however, that in a recent study Langousis and Koutsoyiannis (2006) developed a stochastic framework capable of reproducing simultaneously the long-term and the short-term stochastic structure of hydrological processes, avoiding the use of disaggregation. While disaggregation and downscaling are similar in nature, they also have a difference that distinguishes them. Downscaling aims at solely producing a precipitation field y with the required statistics at the scale of interest, being statistically consistent with the given field x at the finer scale. Disaggregation demands full and precise consistency, which introduces an equality constraint in the problem of the form

Cy¼x

ð55Þ

where C is a matrix of coefficients. For example, assuming that x is an annual amount of precipitation at a station and y is the vector consisting of the 12 monthly precipitation values at the

same station, C will be a row vector with all its elements equal to 1, so that Equation (55) represents the requirement that the sum of all monthly precipitation amounts must equal the annual amount. Task 1 could be accomplished by running a second meteorological model at the limited area of interest. Such models, known as limited-area models, can have much higher resolution than global models. The description of this type of downscaling, known as dynamical downscaling, because it is based on the atmospheric dynamics, is not within the scope of this chapter. In contrast, a stochastic procedure need not refer to the dynamics, and is generic and appropriate for both downscaling and disaggregation and for all above tasks 1–4. This generic procedure resembles the interpolation procedure described in Section 2.02.3.1.3, but there are two important differences. First, it is necessary to include the error terms in the generation procedure (recall that in interpolation, which is a point estimation, knowing only the mean and variance of the error was sufficient). Second, the generated values y at the different points should be statistically consistent to each other. This precludes the separate application of an algorithm at each point of interest and demands simultaneous generation at all points. In turn, this demands that the error terms in different points should be correlated to each other. All these requirements could be summarized in the linear generation scheme

y ¼A xþB v

ð56Þ

where A and B are matrixes of coefficients and v is a vector of independent random variables, so that the term Bv ¼: e corresponds to the error term in interpolation (cf. Equation (29)). In disaggregation, Equation (56) should be considered simultaneously with Equation (55). For Gaussian random fields without intermittency, the application of Equations (55) and (56) is rather trivial. However, the intermittency of the rainfall processes and the much-skewed distributions at fine timescales are severe obstacles for rainfall disaggregation. To overcome such obstacles, several researchers have developed a plethora of rather ad hoc disaggregation models (see review by Koutsoyiannis (2003b)). However, the application of the above theoretically consistent scheme is still possible, if combined with a stochastic model, accounting for intermittency (e.g., a Bartlett– Lewis model), and if an appropriate strategy is used to implement Equation (55). Such a strategy includes recursive application of Equation (56) until the error in Equation (55) becomes relatively low, and is followed by correction of the error of the accepted final iteration by appropriate adjusting procedures, which should not alter the covariance structure of the precipitation field. The general strategy of stochastic disaggregation is described in Koutsoyiannis (2001) and two implementations for temporal rainfall disaggregation at a fine (hourly) scale at a single site and at multiple sites are described in Koutsoyiannis and Onof (2001) and Koutsoyiannis et al. (2003), respectively. The models described in the latter two papers, named Hyetos and MuDRain, respectively, are available online, and have been used in several applications worldwide. Typical results of the two models are shown in Figures 25 and 26, respectively.

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Figure 25 Typical screens produced by the Hyetos software during disaggregation of daily to hourly rainfall data, where plots in green and red refer to disaggregated and original data respectively. Upper-left panel shows typical hyetographs, where the green (disaggregated) plot is the result of the storms shown in Figure 23 converted to a hyetograph at an hourly scale. Notice that while daily totals match, the temporal distribution of rainfall differs in the disaggregated and original hyetographs. However, in the statistical sense, the disaggregated series resembles the original, as shown in the other panels comparing statistics of disaggregated and original series.

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Figure 26 While, as shown in Figure 25 (upper-left panel), in single-variate disaggregation, the produced hyetographs resemble the actual ones only in a statistical sense, multivariate disaggregation reproduces the actual shapes of hyetographs provided that fine-scale (e.g., hourly) data exist in at least one of the stations. The two panels show a comparison of historical (marked H) and simulated (by the MuDRain disaggregation model; marked S) hyetographs on a day with relatively high rainfall (B16 mm) at two rain gauges (2 and 5) in the Brue catchment located in South-Western England. From Koutsoyiannis D, Onof C, and Wheater HS (2003) Multivariate rainfall disaggregation at a fine timescale. Water Resources Research 39(7): 1173 (doi:10.1029/2002WR001600).

2.02.4.5 Multifractal Models Rainfall models of multifractal type have for a long time been known to accurately reproduce several statistical properties of actual rainfall fields in finite but practically important ranges

of scales: typically from below 1 h to several days in time and from below 10 km to more than 100 km in space (Schertzer and Lovejoy, 1987, 1989; Tessier et al., 1993; Fraedrich and Larnder, 1993; Olsson, 1995; Lovejoy and Schertzer, 1995; Over and Gupta, 1996; Carvalho et al., 2002; Nykanen and

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Harris, 2003; Kundu and Bell, 2003; Deidda et al., 2004, 2006; Gebremichael and Krajewski, 2004; Calenda et al., 2005; Gebremichael et al., 2006; Veneziano and Langousis, 2005; Garcı´a-Marı´n et al., 2007; Langousis and Veneziano, 2007). These properties include the scaling of the moments of different orders (Schertzer and Lovejoy, 1987; Menabde et al., 1997; Deidda et al., 1999; Deidda, 2000), the power law behavior of spatial and temporal spectral densities (Olsson, 1995; Tessier et al., 1996; Deidda et al., 2004, 2006), the alteration of wet and dry intervals (Over and Gupta, 1996; Schmitt et al., 1998; Olsson, 1998; Gu¨ntner et al., 2001; Langousis and Veneziano, 2007), and the distribution of extremes (Hubert et al., 1998; Veneziano and Furcolo, 2002; Veneziano and Langousis, 2005; Langousis and Veneziano, 2007; Langousis et al., 2007; Veneziano et al., 2009). Significant deviations of rainfall from multifractal scale invariance have also been pointed out. These deviations include breaks in the power-law behavior of the spectral density (Fraedrich and Larnder, 1993; Olsson, 1995; Menabde et al., 1997), lack of scaling of the non-rainy intervals in time series (Schmitt et al., 1998), differences in scaling during the intense and moderate phases of rainstorms (Venugopal et al., 2006), the power deficit at high frequencies relative to multifractal models (Perica and Foufoula-Georgiou, 1996a, 1996b; Menabde et al., 1997; Menabde and Sivapalan, 2000), and more complex deviations as described in Veneziano et al. (2006a). Next, we review some basic properties of stationary multifractal processes and discuss a simple procedure to construct discrete multifractal fields based on the concept of multiplicative cascades. For a detailed review on the generation of multifractal processes and their applications in hydrological modeling and forecasting, the reader is referred to Veneziano and Langousis (2010). Let i ðdÞ ðtÞ be the average rainfall intensity averaged over timescale d at time t. The stochastic process i ðdÞ ðtÞ is said to be stationary multifractal if, for any timescale d, its statistics remain unchanged when the time axis is contracted by a factor r41 and the intensity is multiplied by a random variable ar , that is, d

i ðd=rÞ ðtÞ ¼ ar iðdÞ ðtÞ d

has zero mean, it can be viewed as a special case of the multifractal process in which the random variable ar is replaced by a deterministic power function of resolution r. A property of stationary multifractal processes, which has been used to verify multifractality, is that the spectral density s(o) behaves like ob where o is the frequency, and bo1 is a constant (e.g., Fraedrich and Larnder, 1993; Olsson, 1995, Deidda et al., 2004; Hsu et al., 2006). More comprehensive checks of multifractality involve the dependence of statistical moments of different orders on scale. In particular, under perfect multifractality E½ði ðdÞ Þ q p E½ðar Þq p dKðqÞ p r KðqÞ, where K(q) is a convex function, usually referred to as moment-scaling function (Gupta and Waymire, 1990; Veneziano, 1999). All concepts and methods are readily extended to space–time rainfall (Veneziano et al., 2006b). A simple procedure to construct discrete stationary multifractal fields is based on iterative application of Equation (57) starting from a large timescale dpdmax and gradually decreasing the timescale (i.e., at resolutions rpmn, where m 4 1 and nX1 are integers). The contraction by the same factor r ¼ m at each step simplifies generation, since only the distribution of ar  am is needed. This forms the concept of socalled isotropic discrete multiplicative cascade. Its construction in the D-dimensional cube SD starts at level 0 with a single tile O01  SD with constant unit intensity inside O01 . At level n ¼ 1, 2, y (or equivalently at resolutions r ¼ mD, m2D, y) each tile at the previous level n  1 is partitioned into mD tiles where m 4 1 is the integer multiplicity of the cascade. The intensity inside each cascade tile Oni (i ¼ 1, y, mnD) is obtained by multiplying that of the parent tile at level n  1 by an independent copy yi of a unit-mean random variable y, called the generator of the cascade. Clearly, for r ¼ mnD, ar ¼ y1 y2 ? yn . For illustration, Figure 27 shows a simulated realization of a 2D binary (i.e., m ¼ 2) discrete multiplicative cascade developed to level n ¼ 8.

2.02.5 Precipitation and Engineering Design 2.02.5.1 Probabilistic versus Deterministic Design Tools

ð57Þ

where ¼ denotes equality in (any finite-dimensional) distribution. The notation implies that the distribution of ar depends only on r and not on time t or the intensity i ðdÞ. Obviously, the mean of ar is 1 and furthermore ar is assumed to be stochastically independent of i ðdÞ at the higher scale d. The distribution of ar characterizes the scaling properties as well as many other characteristics of the rainfall process including the marginal distribution, intermittency, distribution of extremes, etc. Equation (57) need not apply for arbitrarily large timescales but rather applies up to a maximum scale d ¼ dmax. In rainfall, dmax seems to be of the order of several days and it is representative of the mean interarrival time of rainfall events (Langousis and Veneziano, 2007; Langousis et al., 2007; Veneziano et al., 2007). We note for comparison that the related equation in the simple scaling (HK) repred sentation of Section 2.02.1.5. is ði ðd=rÞ  mÞ ¼ r 1H ði ðdÞ  mÞ or 1H 1H ðdÞ ðd=rÞ d i ¼ mð1  r Þþr i , so that, when the HK process

The design and management of flood protection works and measures require reliable estimation of flood probability and risk. A solid empirical basis for this estimation can be offered by flow-observation records with an appropriate length, sufficient to include a sample of representative floods. In practice, however, flow measurements are never enough to support flood modeling. The obvious alternative is the use of hydrologic models with rainfall input data to generate streamflow. Notably, even when flow records exist, rainfall probability still has a major role in hydrological practice; for instance, in major hydraulic structures, the design floods are estimated from appropriately synthesized design storms (e.g., US Department of the Interior, Bureau of Reclamation, 1977, 1987; Sutcliffe, 1978). The need to use rainfall data as the basis of hydrologic design becomes even more evident in the study of engineering structures and urban water-management systems that modify the natural environment, so that past flood records, even if they exist, are no longer representative of the future modified system.

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69

Figure 27 Simulated realization of a 2D stationary multifractal field. The random variable y is taken to be lognormal with unit mean value and logvariance ðslny Þ2 ¼ 0:2lnð2Þ.

Hydrologic design does not necessarily require full modeling of the rainfall process, of the type discussed in Section 2.02.4. Usually, in design studies, the focus is on extreme rainfall, which, notably, may not be represented well in such models, which are better for the average behavior of rainfall. However, historically, the perception of intense rainfall and the methodologies devised to model it have suffered from several fallacies spanning from philosophical to practical issues, which we describe next to cast a warning against their acceptance and use. The first fallacy is of a rather philosophical type. As discussed in Section 2.02.1.5, the modeling of the rainfall process in pure deterministic terms has been proven to be problematic. However, deterministic thinking in science is strong enough, so that after the failure in providing full descriptions, it was headed to determining physical bounds to precipitation in an attempt to design risk-free constructions or practices. The resulting concept of probable maximum precipitation (PMP), that is, an upper bound of precipitation that is physically feasible (World Meteorological Organization, 1986), is perhaps one of the biggest failures in hydrology. Using elementary logic, we easily understand that even the terminology is self-contradictory, and thus not scientific. Namely, the word probable contradicts the existence of a deterministic limit. Several methods to determine PMP exist in literature and are described in World Meteorological Organization (1986). However, examination in depth of each of the specific methods separately will reveal that they are all affected by logical inconsistencies. While they are all based on the assumption of the existence of a deterministic upper limit, they determine this limit statistically. This is obvious in the so-called statistical approach by Hershfield (1961, 1965), who used 95 000 station-years of annual maximum daily rainfall belonging to 2645 stations, standardized each record, and found the maximum over the 95 000 standardized values. Naturally, one of the 95 000 standardized values would be the greatest of all others, but this is not a deterministic limit to call PMP (Koutsoyiannis, 1999). If one examined 95 000 additional measurements, one might have found an even higher value. Thus, the logical problem here is the incorrect interpretation that an observed maximum in precipitation is a physical upper limit.

The situation is perhaps even worse with the so-called moisture maximization approach of PMP estimation (World Meteorological Organization, 1986), which seemingly is more physically (hydrometeorologically) based than the statistical approach of Hershfield. In fact, however, it suffers twice by the incorrect interpretation that an observed maximum is a physical upper limit. It uses a record of observed dew point temperatures to determine an upper limit, which is the maximum observed value. Then it uses this limit for the so-called maximization of an observed sample of storms, and asserts the largest value among them as PMP. Clearly, this is a questionable statistical approach, because (1) it does not assign any probability to the value determined and (2) it is based only on one observed value (known in statistics as the highest-order statistic), rather than on the whole sample, and thus it is enormously sensitive to one particular observation of the entire sample (Papalexiou and Koutsoyiannis, 2006; Koutsoyiannis, 2007). Thus, not only does the determination of PMP use a statistical approach (rather than deterministic physics), but it uses bad statistics. The arbitrary assumptions of the approach extend beyond the confusion of maximum observed quantities with physical limits. For example, the logic of moisture maximization at a particular location is unsupported given that a large storm at this location depends on the convergence of atmospheric moisture from much greater areas. Rational thinking and fundamental philosophical and scientific principles can help identify and dispel such fallacies. In particular, the Aristotelian notions of potentia (Greek, dynamis) and of potential infinite (Greek, apeiron; Aristotle, Physics, 3.7, 206b16) that ‘‘exists in no other way, but ... potentially or by reduction’’ (and is different from mathematical complete infinite) would help us to avoid the PMP concept. In fact, this does not need a great deal of philosophical penetration. The same thing is more practically expressed as ‘‘conceptually, we can always imagine that a few more molecules of water could fall beyond any specified limit’’ (Dingman, 1994). Yet, the linkage to the Aristotelian notions of potentia and potential infinity may make us more sensitive in seeing the logical inconsistencies (see also Koutsoyiannis, 2007). According to Popper (1982) the extension of the Aristotelian idea of potentia in modern terms is the notion of

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probability. Indeed, probability provides a different way to perceive the intense rainfall and flood and to assign to each value a certain probability of exceedance (see next session) avoiding the delusion of an upper bound of precipitation and the fooling of decision makers that they can build risk-free constructions. In this respect, the criticism of the PMP and the probable maximum flood (PMF) involves logical, technical, philosophical, and ethical issues (e.g., Benson, 1973). One typical argument against the use of probabilistic approaches, in favor of PMP, which is very old yet popular even today, has been stated by Horton (1931; from Klemes, 2000), ‘‘It is, however, important to recognize the nature of the physical processes involved and their limitations in connection with the use of statistical methods. y Rock Creek cannot produce a Mississippi River flood any more than a barnyard fowl can lay an ostrich egg.’’ However, this argument reveals an incorrect perception of probability and statistics. In a probability theoretic context, there is not a logical inconsistency. Assuming, for example, that the annual peak flood of the Mississippi river (xM) is on the average (mM), a million times larger than the average (mC) flood of a certain small creek (xC), and assuming that both xM and xC have a lognormal distribution with standard deviation slnx of logarithms of about 0.3 (which is roughly equal to the coefficient of variation of the annual flood peaks, assumed equal in the two streams), one can readily find that the probability that the flood in the creek xC in some year exceeds the mean annual flood mM of Mississippi is F*(z): ¼ 1  FG(z) where FG is the standard normal distribution function and z ¼ ln(mM/mC)/slnx or z ¼ ln(106)/0.3 ¼ 46. For large z, the approximation ln F*(z) ¼  (1/2)[ln(2pz2) þ z2] holds (e.g., Abramowitz and Stegun, 1965); hence ln F*(z) ¼  1062.75, so that the probability of exceedance is F*(z) ¼ 10462. That is, according to the probabilistic approach, the return period of the event that the small creek flood matches or exceeds the mean annual flood of the Mississippi is 10462 years. Assuming that the age of the universe is of the order of 1010 years, one would wait, on the average, 10452 times the age of the universe to see this event happen – if one foolishly hoped that the creek, the Mississippi, and the Earth would exist for such a long time. Evidently, such a low probability could be regarded as synonymous to impossibility, which shows that the probabilistic approach does not regard the floods of Mississippi equivalent to those of a small creek (see also an example about the age of a person by Feller (1950)).

2.02.5.2 Extreme Rainfall Distribution Having been exempted from the concept of an upper limit to precipitation and having adopted a probabilistic approach, the real problem is how the rainfall intensity grows as the probability of exceedance decreases. Clearly, as the probability of exceedance tends to zero, the intensity tends to infinity. There exists a mathematically proven lower limit to the rate of this growth, which is represented by an exponential decay of the probability of exceedance with intensity. The alternative is a power-low decay and, as already mentioned in Section 2.02.1.5, the two options may lead to substantial differences in design quantities for high return periods. In this respect, the

most important questions, which have not received definite answers yet, are again related to the notion of infinity. Accordingly, the distribution tails are important to know in engineering design. However, the study of the tails is difficult and uncertain because the tails refer to infrequent events that require very long records to appear. Traditionally, rainfall records are analyzed in two ways. The most frequent is to choose the highest of all recorded precipitation intensities (for a given averaging timescale) at each year and form a statistical sample with size equal to the number of years of the record. The other is to form a sample with all recorded intensities over a certain threshold irrespective of the year they occurred. Usually, the threshold is chosen high enough, so that the sample size is again equal to the number of years of the record. This however is not necessary: it can well be set equal to zero, so that all recorded intensities are included in the sample. However, the threshold simplifies the study and helps focus the attention on the distribution tail. If x1 ; x2 ; y; xn are random variables representing the recorded average intensities within a year at nonoverlapping time periods equal to a chosen timescale d, then the maximum among them y :¼ maxðx1 ; x2 ; y; xn Þ has a distribution function Hn(y) fully dependent on the joint distribution function of xi . Assuming that xi are IID with common distribution function F(x), then Hn(x) ¼ [F(x)]n. If n is not constant, but rather can be regarded as a realization of a random variable (corresponding to the fact that the number of rainfall events is not the same in each year) with Poisson distribution with mean n, then the distribution function H becomes (e.g., Todorovic and Zelenhasic, 1970; Rossi et al., 1984)

HðxÞ ¼ expfn½1  FðxÞg

ð58Þ

In particular, if the threshold has been chosen with the above rule (to make the sample size equal to the number of years of the record) then obviously n ¼ 1. Equation (58) expresses in a satisfactory approximation, the relationship between the above two methodologies and the respective distributions F and H. The two options discussed above are then represented as follows: 1. Exponential tail.

FðxÞ ¼ 1  expðx=l þ cÞ; HðxÞ ¼ exp½expðx=l þ cÞ; x  lc

ð59Þ

where l 4 0 and c 4 0 are parameters, so that lc represents the specified threshold. Here F is the exponential distribution and H is the Gumbel distribution, also known as extreme value type I (EV1) distribution. 2. Power tail.

x  FðxÞ ¼ 1½1 þ k  c 1=k ;  h l  i1=k  x ; HðxÞ ¼ exp  1 þ k  c l

x  lc

ð60Þ

where l 4 0, c 4 0 and k 4 0 are parameters, and lc represents the specified threshold. Here F is the generalized Pareto distribution (a generalized form of Equation (18)) and H is the generalized extreme value (GEV) distribution.

Precipitation

In the case k 4 0 considered here, GEV is also called the extreme value type II (EV2) distribution. The case ko0 is mathematically possible and is called the extreme value type III (EV3) distribution. However, this is inappropriate for rainfall as it puts an upper bound (lc) for x, which is inconsistent. The case k ¼ 0, corresponds precisely to the exponential tail (exponential and Gumbel distributions). For years, the exponential tail and the Gumbel distribution have been the prevailing models for rainfall extremes, despite the fact that they yield unsafe (the smallest possible) design rainfall values. Recently, however, their appropriateness for rainfall has been questioned. Koutsoyiannis (2004a, 2005a, 2007) discussed several theoretical reasons that favor the power/EV2 over the exponential/EV1 case. As already mentioned (Section 2.02.1.5.5), Koutsoyiannis (2004b, 2005a) compiled an ensemble of annual maximum daily rainfall series from 169 stations in the Northern Hemisphere (28 from Europe and 141 from the USA) roughly belonging to six major climatic zones and all having lengths from 100–154 years. The analysis provides sufficient support for the general applicability of the EV2 distribution model worldwide. Furthermore, the ensemble of all samples was analyzed in combination and it was found that several dimensionless statistics are virtually constant worldwide, except for an error that can be attributed to a pure statistical sampling effect. This enabled the formation of a compound series of annual maxima, after standardization by the mean, for all stations (see Figure 13, which shows the distribution of a compound sample over threshold of all stations, except one in which only annual maxima existed). The findings support the estimation of a unique k for all stations, which was found to be 0.15. Additional empirical evidence with the same conclusions is provided by the Hershfield’s (1961) data set, which was the basis of the formulation of Hershfield’s PMP method. Koutsoyiannis (1999) showed that this data set does not support the hypothesis of an upper bound in precipitation, that is, PMP. Rather, it is consistent with the EV2 distribution with k ¼ 0.13, while the value k ¼ 0.15 can be acceptable for that data set too (Koutsoyiannis, 2004b). This enhances the trust that an EV2 distribution with k ¼ 0.15 can be regarded as a generalized model appropriate for mid-latitude areas of the Northern Hemisphere. In a recent study, Veneziano et al. (2009) used multifractal analysis to show that the annual rainfall maximum for timescale d can be approximated by a GEV distribution and that typical values of k lie in the range 0.09–0.15 with the larger values being associated with more arid climates. This range of values agrees well with the findings of Koutsoyiannis (1999, 2004b, 2005a). Similar results were provided by Chaouche (2001) and Chaouche et al. (2002). Chaouche (2001) explored a database of 200 rainfall series of various time steps (month, day, hour, and minute) from the five continents, each including more than 100 years of data. Using multifractal analyses, it was found that (1) an EV2/Pareto type law describes the rainfall amounts for large return periods; (2) the exponent of this law is scale invariant over scales greater than an hour (as stated in Section 2.02.4.2, it cannot be otherwise

71

because this is dictated by theoretical reasons); and (3) this exponent is almost space invariant. Other studies have also expressed skepticism for the appropriateness of the Gumbel distribution for the case of rainfall extremes and suggested hyper-exponential tail behavior. Coles et al. (2003) and Coles and Pericchi (2003) concluded that inference based on the Gumbel model for annual maxima may result in unrealistically high return periods for certain observed events and suggested a number of modifications to standard methods, among which is the replacement of the Gumbel model with the GEV model. Mora et al. (2005) confirmed that rainfall in Marseille (a rain gauge included in the study by Koutsoyiannis (2004b)) shows hyper-exponential tail behavior. They also provided two regional studies in the Languedoc-Roussillon region (south of France) with 15 and 23 gauges, for which they found that a similar distribution with hyper-exponential tail could be fitted. This finding, when compared to previous estimations, leads to a significant increase in the depth of rare rainfall. On the same lines, Bacro and Chaouche (2006) showed that the distribution of extreme daily rainfall at Marseille is not in the Gumbel-law domain. Sisson et al. (2006) highlighted the fact that standard Gumbel analyses routinely assign near-zero probability to subsequently observed disasters, and that for San Juan, Puerto Rico, standard 100-year predicted rainfall estimates may be routinely underestimated by a factor of two. Schaefer et al. (2006) using the methodology by Hosking and Wallis (1997) for regional precipitation-frequency analysis and spatial mapping for 24-h and 2-h durations for the Washington State, USA, found that the distribution of rainfall maxima in this State generally follows the EV2 distribution type.

2.02.5.3 Ombrian Relationships One of the major tools in hydrologic design is the ombrian relationship, more widely known by the misnomer rainfall intensity-duration-frequency (IDF) curve. An ombrian relationship (from the Greek ombros, rainfall) is a mathematical relationship estimating the average rainfall intensity i over a given timescale d (sometimes incorrectly referred to as duration) for a given return period T (also commonly referred to as frequency, although frequency is generally understood as reciprocal to period). Several forms of ombrian relationships are found in the literature, most of which have been empirically derived and validated by the long use in hydrologic practice. Attempts to give them a theoretical basis have often used inappropriate assumptions and resulted in oversimplified relationships that are not good for engineering studies. In fact, an ombrian relationship is none other than a family of distribution functions of rainfall intensity for multiple timescales. This is because, the return period is tied to the distribution function, that is, T ¼ d/[1  F(x)], where d is the mean interarrival time of an event that is represented by the variable x, typically 1 year. Thus, a distribution function such as one of those described in Section 2.02.4.2, is at the same time an ombrian relationship. This has been made clear in Koutsoyiannis et al. (1998) who showed that the empirical considerations usually involved in the construction of

72

Precipitation

ombrian curves are not necessary at all, and create difficulties and confusion. However, the direct use in engineering design of a fully consistent multiscale distribution function may be too complicated. Simplifications are possible to provide satisfactory approximations, given that only the distribution tail is of interest and that the range of scales of interest in engineering studies is relatively narrow. Such simplifications, which were tested recently and were found to be reasonable (Papalexiou and Koutsoyiannis, 2009) are 1. The separability assumption, according to which the influences of return period and timescale are separable (Koutsoyiannis et al., 1998), that is,

iðd; TÞ ¼

aðTÞ bðdÞ

ð61Þ

where a(T) and b(d) are mathematical expressions to be determined. 2. The use of the Pareto distribution for the rainfall intensity over some threshold at any timescale, as discussed in Section 2.02.5.2; this readily provides a simple expression for a(T). 3. The expression of b(d) in the simple form

bðdÞ ¼ ð1 þ d=yÞ Z

ð62Þ

where y40 and Z40 are parameters. A justification of this relationship, which is a satisfactory approximation for timescales up to a few days, can be found in Koutsoyiannis (2006a). Based on assumptions 1–3, we easily deduce that the final form of the ombrian relationship is

iðd; TÞ ¼ l0

ðT=dÞ k  c0 ð1 þ d=yÞ Z

ð63Þ

where c0 40, l0 40 and k40 are parameters. In particular, as discussed in Section 2.02.5.2, k is the tail-determining parameter and unless a long record exists, which could support a different value, it should be assumed k ¼ 0.15. Equation (63) is dimensionally consistent, if y has units of time (as well as d), l0 has units of intensity, and k and c are dimensionless. The numerator of Equation (63) differs from a pure power law that has been commonly used in engineering practice, as well as in some multifractal analyses. Consistent parameter-estimation techniques for ombrian relationships have been discussed in Koutsoyiannis et al. (1998) as well as in Chapter 2.18 Statistical Hydrology.

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2.03 Evaporation in the Global Hydrological Cycle AJ Dolman and JH Gash, VU University Amsterdam, Amsterdam, The Netherlands & 2011 Elsevier B.V. All rights reserved.

2.03.1 2.03.2 2.03.2.1 2.03.2.2 2.03.2.3 2.03.3 2.03.4 2.03.5 References

Introduction General Theory of Evaporation Vegetated Surfaces Bare Soil Open Water and Lakes Regional and Equilibrium Evaporation Trends and Variability in Global Evaporation Summary and Conclusions

2.03.1 Introduction Evaporation is the transfer of moisture from a particular surface to the overlying atmosphere. The physical process of evaporation consists of the exchange of water molecules between a free water surface and the air. The surface can be any among the following nonexhaustive list: a lake, the inside of plant leaves, the water surface adhering to a soil conglomerate, or the surface of a soil or canopy during or just after rain. The evaporation rate is expressed as the quantity of water evaporated per unit area per unit time from a (water) surface under existing atmospheric conditions. This chapter describes the progress in understanding evaporation at the local scale, both from an observational and a conceptual perspective. It then moves on to the global scale, through a discussion of regional scale feedbacks. The main conclusion is that while we have gained considerable understanding in local scale evaporation, the data sets required to study the impact of evaporation on the global water cycle are still lacking. Evaporation at the Earth’s surface is constrained both by the energy available to convert liquid water into vapor and by the capability of the surrounding air to transfer moisture away from the saturated surface. At the surface, these constraints are best expressed through the energy balance equation and the transfer equations of latent and sensible heat:

Rg ð1  zÞ þ Lk  Lm ¼ lE þ H þ G þ dS

ð1Þ

lE ¼ rlKv

qq qz

ð2Þ

H ¼ rcp Kh

qT qz

ð3Þ

b¼

H lE

ð4Þ

with Rg the incoming short-wave radiation (W m2), z the short-wave albedo, Lk and Lm the incoming and outgoing long-wave radiation (W m2), lE the evaporation (or latent heat flux, W m2) (with l the latent heat of vaporization and E the mass flux), H the sensible heat flux (W m2), G the soil heat flux (W m2), dS the change in heat storage in the

79 80 81 82 82 83 83 85 85

biomass and atmosphere below a reference height above the surface (W m2), Kv and Kh the transfer coefficients for water vapor and heat, respectively, q q/q z the vertical gradient in specific humidity (g kg1), and similarly q T/q z is the vertical gradient in temperature (K), r is the density of air (kg m3), and cp the specific heat of air (J kg1 K1). b is the Bowen ratio and a useful indicator of the dryness of the surface because it shows how the available energy is partitioned into sensible and latent heat. Factors influencing the rate of evaporation can easily be determined from the above equations: radiation as the limit of available energy, the gradient in moisture, or the specific humidity deficit and factors that influence the transfer coefficients such as wind speed and roughness of the surface, and, crucially, when dealing with evaporation from leaves (transpiration), water availability (soil moisture). Thus, despite the complexity of the interaction between a partially wet surface and the atmosphere, it may be possible to simplify this interaction and to approximate evaporation by considering only two factors, available energy (radiation) and water availability. Available energy determines the maximum evaporation possible for the given climatic conditions and unlimited water availability, that is, potential evaporation (see also Allen et al., 1998). Water availability can be characterized by the amount of precipitation. Under dry conditions, potential evaporation exceeds precipitation, and the actual evaporation from an area will approach the amount of precipitation received. Conversely, under wet conditions, water availability exceeds potential evaporation and actual evaporation will approach asymptotically potential evaporation. This relationship was first used by Budyko (1974) to set a constraint on global evaporation. Later, Milly and Dunne (2002) suggested that the long-term water balance of a catchment is determined by the interaction of supply (precipitation) and demand (potential evaporation), mediated by soil moisture storage. At longer timescales, the change in soil moisture can be neglected. Figure 1 shows results of Zhang et al. (2001), from a simple model based on the above considerations for grass and forested catchments, using catchment discharge and precipitation data, that allow evaporation to be calculated as the residual of these two. These results suggest that at the scale of catchments such a simple approach works remarkably well, with a linear (1:1) relation between rainfall and evaporation for catchments that receive up to 500 mm yr1, and an

79

80

Evaporation in the Global Hydrological Cycle 1600

Annual evapotranspiration (mm)

Forest Mixed veg. Pasture 1200

800

400

0

0

500

1000

1500

2000

2500

3000

3500

Annual rainf all (mm) Figure 1 Relationship between annual precipitation and evaporation. The dashed line represents the best fit of a theoretical model (Zhang et al., 2001) for grass-only catchments, the solid line the best fit for forests. From Zhang et al. (2001). (Note that we use the word evapotranspiration here in this graph consistent with its origin. This rather loosely defined term refers to the total of (wet canopy and bare soil) evaporation and transpiration. Throughout the text, we prefer to use the word evaporation as the physical term denoting the process of transforming liquid water into its vapor form.)

asymptotic relation (driven by available energy and not by water availability) for catchments receiving more than 500 mm yr1. The difference between forest and grassland is a result of the high rate of evaporation of intercepted rainfall from forest; interception loss creates a considerable additional evaporative loss for forest (see also Chapter 2.04 Interception). Total evaporation from forest saturates at about 1400 mm yr1, while for grasslands this value is around 900 mm yr1. Direct observations of evaporation have been made since the mid-1990s using the eddy-covariance technique in a global network, Fluxnet (see Aubinet et al., 2001). While there are other techniques based on scintillometry available, these have not been used extensively in networks such as Fluxnet. Evaporation measured by micrometeorological techniques usually refers to dry-canopy evaporation only (transpiration) and contains little information on evaporation during wet-canopy conditions (interception). Consequently, total evaporation values such as those given by Law et al. (2002) should be treated with caution, when interception losses are not explicitly treated. With this caveat in mind, the annual evaporation from Fluxnet data for coniferous forests is 397 (731) mm yr1; for mixed evergreen and deciduous forest, 386 (718) mm yr1; for deciduous broadleaf, 512 (769) mm yr1; for grassland, 494 (7104) mm yr1; and for crops, 666 (767) mm yr1. Grassland and crops have higher dry-canopy evaporation than forest, because they are less strongly coupled to the atmosphere and thus show less stomatal control (e.g., Shuttleworth and Calder, 1979). This also gives forest the possibility to survive occasional drought that would kill off annual grassland species. The Fluxnet data can be used to identify the main controls on evaporation. Wilson et al. (2002) showed how the

partitioning of energy, as expressed by the ratio of sensible heat to latent heat, the Bowen ratio (b), can be used to classify different vegetation types in climate space. Figure 2 shows the position of the individual sites with respect to the magnitude of the latent heat (evaporation) and sensible heat fluxes. In contrast to the annual rates discussed earlier, these data refer to the growing season only. Also shown are lines of constant available energy (the sum of latent and sensible heat) and lines of constant b. Moving from the lower left corner of the diagram to the upper right, the available energy increases, whereas moving from the lower right to the upper left, the value of b increases. Low evaporation rates are found in Sitka spruce, tundra, and boreal forest, and high evaporation rates in deciduous forests and agriculture. Most of the forests have high b’s, implying that much of the energy received is transmitted back to the atmosphere as sensible heat. Deciduous forests tend to have lower b’s with higher evaporation rates than coniferous forests. The average b at tundra sites appears to be close to 1.

2.03.2 General Theory of Evaporation Penman (1948) was among the first to achieve the crucial combination of the energy balance equation (1) with the transfer Equations (2) and (3) to derive an expression for actual evaporation from vegetation well supplied with water. Although the vertical gradients in Equations (2) and (3) could be derived from the differences between air and surface values, measurements of surface temperature are difficult and not made routinely. Penman overcame this problem by introducing the slope of the saturated vapor pressure versus temperature curve, D, approximated as a linear function and

Evaporation in the Global Hydrological Cycle 15

81

=3 T1

=2

T2

12

M

Daily sensible heat flux (MJ m−2 d−1)

ed

ite

rra

ne

an

m

S1

9

=1

cli

at

es

Boreal Canadian

S2 N4 R 2 R1 N1 Y2 N3

I1

6 Sitka spruce

Q3

 = 0.5

M1 P1 T3

O1

N2

K1 G2

K2 F3 G1Y1F1 Q1 H1 A8 L1 F2 Tundra A7A1 M2 I2 D1 L2 A1 A4 X1 A5 B B1 A2 H 3 V1V 2 B4 G3 A6 B5 B3 U 3 G2 A3 J1 Z1 B2 Conifers U2 B6 C1 O2

Q2

HW1 2

X2

3

E2

 = 0.25

Agriculture E1

Deciduous forests

0 0

3

6

9

12 −2

Daily latent heat flux (MJ m

15

d−1)

Figure 2 The daily cumulative sensible heat flux vs. the daily cumulative latent heat flux between days 165 and 235 for the Fluxnet sites analyzed by Wilson et al. (2002). The letter and number codes refer to the sites as given by Wilson et al. (2002). Also shown are lines of constant Bowen ratio (dashed lines) and lines of constant total turbulent energy fluxes (solid diagonal lines). Enclosed areas denote subjective delineations between different vegetation types and climates. From Wilson et al. (2002).

evaluated at air temperature:

D¼

e  ðTs Þ  e ðTa Þ Ts  Ta

ð5Þ

where e* (Ts) is the saturated vapor pressure at surface temperature Ts (K), and e* (Ta) is the saturated vapor pressure at air temperature Ta. The evaporation, E, in mm d1 is given by

E¼

DQ þ gEa Dþg

ð6aÞ

where Q is water equivalent of the net radiation (from the lefthand side of Equation (1)), g the psychrometric constant and

Ea ¼ f ðUÞðe ðTa Þ  ea Þ

ð6bÞ

Ea is the aerodynamic or demand term for evaporation, with f(U) a wind function, and (e* (Ta)  ea) the vapor pressure deficit. The original Penman equation contains an empirical wind function that replaces the transfer coefficients in Equations (2) and (3). This wind function was difficult to generalize, and subsequently Thom and Oliver (1977) provided a more

physical basis including considerations of aerodynamic transfer over rough surfaces by introducing an explicit aerodynamic resistance. Penman applied his equation to bare soil evaporation, vegetated or cropped surfaces, and open water bodies. The equation is still widely used to calculate evaporation from well-watered, short vegetation.

2.03.2.1 Vegetated Surfaces Monteith (1965) (see also Gash and Shuttleworth, 2007) introduced the control of vegetation on evaporation by including a canopy scale resistance rs. This resistance represents the restriction on the transfer of water from the collective saturated surfaces inside the plant stomatal cavities to the air outside the leaves. The resulting equation, now carrying both an aerodynamic (to replace the Penman wind function) and a canopy resistance, is known as the Penman–Monteith (PM) equation and is arguably still the most elegant yet advanced resistance model of evaporation used in hydrological practice today (Shuttleworth, 1993). For a mathematically precise and exact definition and brief historical overview of its development, the reader is referred to Raupach (2001). Most commonly, the PM equation is expressed in specific humidity units (rather than vapor pressure deficit as in Equation (6)) with

82

Evaporation in the Global Hydrological Cycle

evaporation expressed in energy units (W m2) and reads as

lE ¼

rcp ðq * ðTa Þ  qa Þ ra   cp rs 1þ Dþ l ra

DðQ *  GÞ þ

ð7Þ

with D the slope of the saturated specific humidity versus temperature curve and Q* the net radiation (W m2). q* (Ta) is the saturated specific humidity of the air (g kg1) at the reference level and qa the specific humidity at the same level. Two important variables appear in this equation that replace the transfer coefficients of the transfer equations of heat and moisture: the aerodynamic and surface resistance, ra and rs (s m1). (Only in the case of a full canopy cover, can the surface resistance be equal to the canopy resistance. For the derivation of the big-leaf version of the PM model as presented here, this is not important. When the canopy cover is not full, and there is bare ground directly in contact with the overlying atmosphere, the surface resistance is not equal to the canopy resistance and approaches the reciprocal sum of the canopy and an assumed soil resistance.) The PM equation assumes that evaporation and sensible heat originate from the same source in the canopy. The main advantage of the PM equation is that the meteorological driving variables, wind speed, specific humidity deficit, and temperature are required only at a single level above the surface, removing the need for the notoriously difficult observation of surface values. The main obstacle for the practical application of this equation is the estimation of the values for aerodynamic and surface resistance. When there is unlimited supply of water, the PM equation can be used to calculate potential evaporation. It can be shown then to collapse to the Penman equation with an aerodynamic resistance rather than a wind function. The PM equation is now the preferred method of estimating crop water requirements as reference crop evaporation (Allen et al., 1998). When the canopy is wet, the surface resistance equals zero and the PM equation can be used to calculate evaporation of intercepted rainfall (see also Chapter 2.04 Interception). Although there is a wide range of empirical evaporation equations of which some are particular cases of the PM equation (see Brutsaert, 1982; Shuttleworth, 1993), the use of the PM equation is widespread because of its clear physical interpretation.

2.03.2.2 Bare Soil Evaporation from bare soil can be a significant component of the water balance, particularly in semi-arid environments (Wallace and Holwill, 1997). Soil evaporation can be described as a two-stage process. The first stage occurs when the available soil moisture is sufficient to meet the atmospheric demand. This occurs immediately after rainfall or irrigation events. Soil evaporation under these conditions equals potential evaporation. Typically, this stage lasts 1–2 days, although in some cases when evaporative demand is low and the soil contains a high amount of clay, this stage may last for up to 5 days. In the second stage, the amount of soil moisture has dropped and soil evaporation is no longer only restricted by evaporative demand but also by availability of moisture. In

these conditions, the change of soil moisture with time can be described as a desorption process with evaporation proportional to the square root of the time since the start of the process:

lEs ¼

1 aðt  t0 Þ1=2 2

ð8Þ

with lEs the soil evaporation, t the time and t0 the time since the start of second stage drying, Ds is a desorptivity (in units of W m2 d1/2 when evaporation is expressed as a heat flux, or in mm d1/2 when evaporation is expressed as water flux). The desorptivity is assumed constant for a particular soil type. It varies from a value of 2.1 for sandy loam with gravel, to a value of 5 mm d1 for a clay loam soil (Kustas et al., 2002). Although the two-stage process describes soil evaporation at diurnal timescales, extension of the theory to (sub) hourly timescales is straightforward (Brutsaert and Chen, 1996; Porte´Agel et al., 2000). The determination of the desorptivity coefficient can be problematic, as can the identification of the switch from stage 1 to 2 (Kustas et al., 2002). The observed dependence of soil evaporation on available soil moisture suggests the feasibility of a resistance approach that incorporates a dependence of soil surface resistance on soil moisture. Mahfouf and Noilhan (1991) review several such formulations. These approaches can be divided into socalled a- and bs-approaches. In the a-approach, the saturated humidity in the soil pore space is adjusted by a factor a that may be related to soil matrix potential and takes into account that, averaged over a certain depth, the evaporation takes place from a nonsaturated surface. In the bs-approach, the humidity in the pore space at the evaporation front is assumed to be saturated, and bs is the ratio of an aerodynamic resistance to the sum of the aerodynamic and soil surface resistance.

2.03.2.3 Open Water and Lakes The key process controlling evaporation from large lakes (or reservoirs) is the absorption of solar energy. This energy is not absorbed at the surface; because water is a semitransparent medium, the solar radiation is absorbed over depth. The rate of change of absorption with depth depends on the turbidity, but may be significant down to several meters below the surface. Solar energy heats the water up during the spring and summer, and this energy is not available for evaporation; but energy released as the water cools in autumn and winter is available and enhances the evaporation. This creates a phase lag between lake evaporation and the annual radiation cycle. While in the tropics this lag will be small, at high latitudes the phase lag may be as much as 5 or 6 months (Blanken et al., 2000) with the rate of change in storage, dS/dt, being the dominant source of energy for evaporation. The energy available for evaporation is given by

A ¼ Q þ

dS þ Aq dt

ð9Þ

where Aq is the rate of net energy advection due to inflow and outflow of water. Net radiation, Q* , is given by Equation (1)

Evaporation in the Global Hydrological Cycle

with long-wave radiation emitted by the lake as

Lm ¼ esT4w

ð10Þ

e is emissivity, s the Stefan–Boltzmann constant, and Tw the surface temperature of the water (K). Q* over the lake can be estimated from measurements over land, but must take account of the different albedo and different surface temperature of the lake. Like the ocean, lake albedo varies strongly with solar elevation (see Finch and Gash, 2002; Finch and Hall, 2005; Payne, 1972). The Penman (1948) equation (Equation (6)) is often used to estimate lake evaporation, as it appears to remove the need for surface temperature measurement; however, this is not the case as water temperature is still needed to calculate the emitted long-wave radiation. Nevertheless, it should give good results if used with the available energy calculated from Equation (9) and measurements made over the lake (see Linacre, 1993). Working in the tropics where the annual cycle in water temperature is small and energy storage can be neglected, Sene et al. (1991) found good agreement between daily estimates made with the Penman equation and eddy-covariance measurements of evaporation. To overcome the lack of water temperature measurements, Finch and Gash (2002) applied a simple numerical, finite difference scheme to calculate a running balance of lake energy storage. A new value of the water temperature required to force energy closure was calculated at each time step. For a well-mixed lake of known depth, the evaporation could then be calculated from land-based, daily meteorological observations of sunshine hours, relative humidity, wind run, and average air temperature. The model gave good agreement with mass-balance measurements of the water loss from a reservoir with no inflow or outflow.

2.03.3 Regional and Equilibrium Evaporation

D Dþ

cp A l

the parameter a. a can be shown to be unity only when the specific humidity deficit in the PM equation is zero, in other words, when advection is negligible. That this is hardly ever the case proves the fact that most empirical values of a for short crops are of the order 1.2–1.3 (e.g., Brutsaert, 1982). For tall crops, Shuttleworth and Calder (1979) showed convincingly that the equilibrium approach is not appropriate because the physiological control of the forest transpiration reduces a below a value of 1 in dry canopy conditions, while in wet canopy conditions large-scale advection and negative sensible heat fluxes (Stewart, 1977) may form an additional supply of energy and force a to be well above the value for short crops. This emphasizes the important point that for tall crops in particular, it is important to estimate dry and wet canopy evaporation separately (see also Figure 1). Thus, at larger scale, atmospheric conditions can override the surface control by exerting a strong feedback on evaporation through the humidity and temperature of the atmospheric boundary layer. A number of concepts have been derived that use the feedback power of the atmosphere to estimate regional-scale evaporation (e.g., Bouchet, 1963; Morton, 1983). Although McNaughton and Jarvis (1991) show that at larger scale the feedback of the atmospheric boundary layer dampens the effects of surface controls – and thus makes precise estimation of the surface resistance less important – the physical basis of the Bouchet and Morton schemes remains doubtful (see de Bruin, 1983; McNaughton and Spriggs, 1989). Nevertheless, de Bruin showed that the feedback of the increasing atmospheric boundary-layer humidity during the day causes the regional surface conductance to vary less than if the feedback were neglected. The relatively large confidence of hydrologists in using potential evaporation formulas probably finds its physical explanation in this feedback.

2.03.4 Trends and Variability in Global Evaporation

Priestley and Taylor (1972) showed that the Bowen ratio would approach a constant value defined by b ¼ s/l when air moves over a moist surface, and gradients of temperature and specific humidity with height are small or become saturated with respect to moisture. Combining this insight with the PM equation, and setting the second term above the nominator to zero as well as the surface or canopy resistance (rs ¼ 0) as would be appropriate for a moist surface, yields the equilibrium evaporation (see also Brutsaert, 1982; Raupach, 2001):

lE ¼

83

ð11Þ

Equilibrium evaporation as defined by these authors refers to the lower limit of evaporation from a moist surface where the specific humidity deficit of the second term in the nominator of the PM equation has become zero as a result of contact of air with a moist surface over a very long fetch. It can easily be shown that the second term in the nominator of the PM equation (Equation (7)) now represents the departure from this equilibrium. Priestley and Taylor (1972) represented this departure from equilibrium evaporation by

Globally, evaporation from the land surface to the atmosphere amounts to 71 103 km3 yr1 (Baumgartner and Reichel, 1975). It is the key return flow in the hydrological cycle from the surface on which the precipitation falls, back to the atmosphere. Evaporation from the oceans is a far larger component of the hydrological cycle at an estimated 428 103 km3 yr1 (Baumgartner and Reichel, 1975). A recent multi-model ensemble of 11 state-of-the-art land surface models (Dirmeyer et al., 2006) estimated annual evaporation over a range of 58 103 to 85 103 km3, indicating the uncertainty in our ability to model evaporation from land. Oki and Kanae (2006) estimated total terrestrial evaporation at 66 103 km3 yr1. Thus, although it is an important component of the hydrological cycle, the exact magnitude and variability, both spatially and temporally, of evaporation from land remains highly uncertain. Figure 3 shows one of the few available estimates of the latitudinal distribution of evaporation in mm yr1 for both ocean and land. This estimate (Baumgartner and Reichel, 1975) is based on the balance between precipitation and runoff on land and a variety of other methods (Peixoto and Oort, 1996). Note the relatively large contribution of Southern

84

Evaporation in the Global Hydrological Cycle

Latittude

0 85 75 65 55 45 35 25 15 5 −5 −15 −25 −35 −45 −55 −65 −75 −85

375

750

1125

1500

Evaporation (mm yr−1)

Figure 3 Global evaporation according to Baumgartner and Reichel (1975) for different 101 latitude bands.

ρRg,E

−1 +1

ρP,E 0

+1

0

−1 Figure 4 Multi-model analysis of controls on yearly evaporation. Correlation between yearly evaporation and global radiation (rRg,E), and precipitation (rP,E), for the period 1986–95. Each color corresponds to a unique combination of rRg,E and rP,E. The gray lines (legend) show the global frequency distribution (Teuling et al., 2009).

Hemisphere latitude bands 20–401 S compared to their northern equivalents. This difference is largely due to the greater area of land in the Northern Hemisphere, which evaporates at a significantly lower rate than the ocean. The largest evaporative flux is found in the humid tropics, mainly as a result of large amounts of precipitation and high solar radiation. Compared to the latitudinal distribution of precipitation, the evaporation is more smoothly distributed with a general tendency of decreasing evaporation when moving poleward. Decreases in radiation, global dimming, have caused a debate about an observed decline in pan measured evaporation (e.g., Roderick and Farquhar, 2002) that would be contrary to expectations for a warming climate. Peterson et al. (1995), using data from a network of pan evaporimeters in the US and the former Soviet Union, found a decrease in pan evaporation between 1950 and 1990. However, Fu et al. (2009) in a more comprehensive analysis suggested that, although many observations across the world indicate a general trend of pan evaporation decreasing over the last 50 years, this trend is not universal. A decrease in evaporation presents a

paradox, as with global warming one would expect an increase due to the larger water holding capacity of the atmosphere through the Clausius–Clapeyron equation. An increase in evaporation would also match an increase in precipitation, although this is regionally very variable. Roderick and Farquhar (2002) explained the paradox by relating the decrease in evaporation to a decrease in solar radiation (global dimming). However, the dimming trend has recently reversed into a brightening trend and thus cannot singularly be held responsible for the decrease in pan evaporation. Teuling et al. (2009) presented an analysis of the major controls on evaporation using an ensemble of land surface models forced off line with meteorological data. They investigate the control of two key drivers on evaporation, incoming solar radiation and soil moisture. Figure 4 shows that Europe, North Africa, and North America are characterized by two evaporation regimes: a humid regime with high correlation with radiation expressed through the correlation coefficient rRg,E, but low correlation with precipitation, P(rP,E); and a more arid regime with high rP,E, but low rRg,E. Because radiation and precipitation tend to be negatively correlated,

Evaporation in the Global Hydrological Cycle

Teuling et al. (2009) concluded that yearly variations in evaporation reflect either variations in Rg or P, but not in both. Central Europe is among the regions with the highest rRg,E correlation, while in more arid regions such as the US Midwest and the Sahara, evaporation correlates only with precipitation. Using data from direct observations of evaporation (Fluxnet; see Baldocchi et al., 2001), Teuling et al. were able to reproduce and validate these modeled patterns quite well. Teuling et al. (2009) made a strong argument for a regional approach to explain some of the evaporation trends. Based on the different sensitivities of the various drivers of evaporation (see Figure 4) and the conclusion of other work that a dimming trend has been reversed, they concluded that scenarios of both decreasing actual evaporation with decreasing pan evaporation in regions with ample supply of water (e.g., central Europe), and of increasing evaporation with decreasing pan evaporation (e.g., the US Midwest) are consistent. Using basin-scale discharge data and precipitation they found that evaporation decreased over Europe during the dimming period and increased later, consistent with the high sensitivity of European evaporation to radiation rather than precipitation. During the dimming period, the positive trend in runoff is induced by reduced evaporation, rather than increased precipitation. After 1983, evaporation derived as the residual of precipitation minus runoff, increased in all central European basins during the brightening phase. These results suggest that evaporation trends follow radiation trends in central Europe. In contrast, in the US Midwest the upward trends in evaporation derived as a catchment residual before 1983 are followed by decreasing trends. These may be explained by trends in precipitation combined with high correlation between solar radiation and evaporation as inferred from Figure 2. Next to the analysis of Teuling et al. (2009) that attribute changes in evaporation to either radiation or water availability (precipitation), recent studies of changes in pan evaporation in Australia attribute most of the reduction in pan evaporation to reduced wind speed (Roderick et al., 2007; Rayner, 2007; McVicar et al., 2008). The cause of such a reduction in regional wind speed is not certain, although wind speed reductions have been widely reported in mid-latitudes in both hemispheres (see also Shuttleworth, 2009).

2.03.5 Summary and Conclusions Evaporation is an important, but regionally a still poorly quantified term in the global water balance. At global scale its determination as a residual of the continental scale water balance hinges on adequate estimation of precipitation and river discharge. Estimates obtained by direct bottom-up modeling vary considerably. This is in some contrast to our understanding of the basic physics of evaporation, that is well known, as for instance is shown by the Penman and PM equations. Application of these equations for use at local scales, for instance irrigation practice, is widespread (e.g., Allen et al., 1998), and to a large extent very successful. The importance of evaporation in the global hydrological cycle critically encompasses two related aspects: its direct role as a term in the water budget, and its potential to impact

85

weather and climate processes, by changing aspects of the surface energy balance and boundary layer. Both these roles depend on the balance between the controlling forces of evaporation, surface moisture and available energy. Where only water availability is limiting and radiation plentiful, such as in the (semi)-arid tropics, evaporation may add moisture to the overlying air that can then tip the balance to produce precipitation. These semi-arid areas have been found (e.g., Koster et al., 2002) to be sensitive to land surface precipitation feedbacks. In these feedbacks the role of evaporation is critical. On the other hand, in areas where both precipitation and radiation are not limiting, large-scale evaporation appears constrained by the available energy. Under these conditions, equations such as the Priestly–Taylor equation that do not explicitly take account of water availability on average perform well. Thus, to be able to predict the effect of, for instance, landuse change on evaporation (and resulting catchment discharge), one would need to determine first which process, if any, is limiting. In cases where water availability is limiting, atmospheric feedbacks may also become important and simple bottom-up estimates of the change in evaporation may be wrong. In cases where neither water nor energy is limiting, to first order a bottom-up estimate based on energy constraints would be appropriate. Changes in the pattern of largescale moisture recycling, such as found in the Amazon (e.g., Meesters et al., 2009), may however be important to estimate changes in the resulting precipitation climate. The debate whether and where evaporation is increasing or decreasing is fundamental to our understanding of the role of evaporation in the global water cycle and climate. Improved, high-quality data sets are needed to provide benchmarks for climate models and to increase our process understanding. Currently, such data sets unfortunately do not exist. It is worth noting that the role of evaporation in climate is not related only to its direct effects on the hydrological cycle. Through the influence evaporation exerts on the partitioning of the energy balance, the effects on climate are also seen in surface temperatures. High evaporation keeps surfaces cool; low evaporation makes them hot. Seneviratne et al. (2006) and Fischer et al. (2007) used regional and global climate modeling to investigate the role of land surface atmosphere feedbacks on temperature in a changing climate. They concluded that soil moisture through its effect in reducing evaporation has a major impact on the variability and mean and maxima of surface temperatures in Europe. Fischer et al. (2007) concluded that land–atmosphere interactions over drought regions account for typically 50–80% of the number of hot days in a Northern Hemisphere summer. This is mainly due to local effects through the limitation of evaporation (and increased sensible heat flux) due to drought conditions. Drought conditions may also have remote effects on areas around or outside the actual drought region, through changes in atmospheric circulation and advection of air masses. These mechanisms can enhance an existing anticyclonic circulation over, or slightly downstream of, a drought anomaly. Evaporation plays a key role in the global water cycle and hence in the global climate system. There have, however, been very few attempts to produce robust estimates of evaporation based on a global approach. Yet such data are urgently needed

86

Evaporation in the Global Hydrological Cycle

to validate and constrain current climate models. Thus, despite that at the practical level, considerable advances have been made in our ability to estimate and observe evaporation at local level, our understanding of evaporation in the global climate system (e.g., Kleidon and Schymanski, 2008) still shows significant gaps. The challenge for the next decade in evaporation research is to fill these gaps.

References Allen RG, Pereira LS, Raes D, and Smith M (1998) Crop Evapotranspiration – Guidelines for Computing Crop Water Requirements, p. 300. Rome: Irrigation and Drainage, FAO. Aubinet M, Chermanne B, Vandenhaute M, Longdoz B, Yernaux M, and Laitat E (2001) Long term carbon dioxide exchange above a mixed forest in the Belgian Ardennes. Agricultural and Forest Meteorology 108: 293--315. Baldocchi D, Falge E, Gu L, et al. (2001) FLUXNET: A new tool to study the temporal and spatial variability of ecosystem-scale carbon dioxide, water vapor, and energy flux densities. Bulletin of the American Meteorological Society 82: 2415--2434. Baumgartner A and Reichel E (1975) Die Weltwasserbilanz, 179pp. Mu¨nchen: Oldenbourg Verlag. Blanken PD, Rouse WR, Culf AD, et al. (2000) Eddy covariance measurements of evaporation from Great Slave Lake, Northwest Territories, Canada. Water Resources Research 36: 1069--1077. Bouchet RJ (1963) Evapotranspiration re´elle et potentielle, signification climatique, Int. Assoc. Sci. Hydrol., Proc. Berkeley, Calif. Symp., Publ. 62: 134–142. Brutsaert W (1982) Evaporation into the Atmosphere. Dordrecht: Kluwer. Brutsaert W and Chen D (1996) Diurnal variation of surface fluxes during thorough drying (or severe drought) of natural prairie. Water Resources Research 32: 2013--2019. Budyko MI (1974) Climate and Life, 508p. New York: Academic Press. De Bruin HAR (1983) A model of the Priestley–Taylor parameter, a. Journal of the Applied Meteorology 22: 572--578. Dirmeyer PA, Gao X, Zhao M, Guo Z, Oki T, and Hanasaki N (2006) GSWP-2: Multimodel analysis and implications for our perception of the land surface. Bulletin of the American Meteorological Society 87: 1381--1397. Finch JW and Gash JHC (2002) Application of a simple finite difference model for estimating evaporation from open water. Journal of Hydrology 255: 253--259. Finch JW and Hall RL (2005) Evaporation from lakes. In: Anderson MG (ed.) Encyclopedia of Hydrological Sciences, pp. 635--646. Chichester: Wiley. Fischer EM, Seneviratne S, Luethi M, and Schaer C (2007) Contribution of land– atmosphere coupling to recent European summer heat waves. Geophysical Research Letters 34: L06707 (doi:10.1029/2006GL029068). Fu G, Charles SP, and Yu J (2009) A critical overview of pan evaporation trends over the last 50 years. Climatic Change 97(1–2): 193--214. Gash JHC and Shuttleworth WJ (2007) Evaporation. Wallingford: IAHS Press. Gedney N, Cox PM, Betts RA, Boucher O, Huntingford C, and Stott PA (2006) Detection of a direct carbon dioxide effect in continental river runoff records. Nature 439: 835--838. Kleidon A and Schymanski S (2008) Thermodynamics and optimality of the water budget on land: A review. Geophysical Research Letters 35: L20404. Koster RD, Dirmeyer PA, Guo Z, et al. (2004) Regions of strong coupling between soil moisture and precipitation. Science 305: 1138–1140. Kustas WP, Prueger JH, and Hipps LE (2002) Impact of using different time-averaged inputs for estimating sensible heat flux of riparian vegetation using radiometric surface temperature. Journal of Applied Meteorology 41: 319--332. Law BE, Falge E, Gu L, et al. (2002) Environmental controls over carbon dioxide and water vapor exchange of terrestrial vegetation. Agricultural and Forest Meteorology 113: 97--120. Linacre ET (1993) Data-sparse estimation of lake evaporation, using a simplified Penman equation. Agricultural and Forest Meteorology 64: 237--256. Mahfouf JF and Noilhan J (1991) Comparative study of various formulations of evaporation from bare soil using in situ data. Journal of Applied Meteorology 30: 1354--1365. McNaughton KG and Jarvis PG (1991) Effects of spatial scale on stomatal control of transpiration. Agricultural and Forest Meteorology 54: 279--302. McNaughton KG and Spriggs TW (1989) An evaluation of the Priestley–Taylor equation. In: Black TA, Spittlehouse DL, Novak MD, and Price DT (eds.) Estimation

of Areal Evaporation, IAHS Publication No. 177, pp. 89–104. Wallingford: IAHS Press. McVicar TR, Van Niel TG, Li LT, et al. (2008) Wind speed climatology and trends for Australia, 1975–2006: Capturing the stilling phenomenon and comparison with near-surface reanalysis output. Geophysical Research Letters 35: L20403 (doi:10.1029/2008GL035627) Meesters AGCA, Dolman AJ, and Bruijnzeel LA (2009) Comment on ‘‘Biotic pump of atmospheric moisture as driver of the hydrological cycle on land’’ by AM Makarieva and VG Gorshkov, Hydrol. Earth Syst. Sci., 11, 1013–1033, 2007. Hydrology and Earth System Sciences 13: 1299--1305. Milly PCD and Dunne KA (2002) Macroscale water fluxes, 2, Water and energy supply control of their interannual variability. Water Resources Research 38(10): 1206 (doi:10.1029/2001WR000760) Monteith JL (1965) Evaporation and environment. In: The State and Movement of Water in Living Organisms. Proceedings of the 19th Symposium Society for Experimental Biology, pp. 205--234. Swansea: Cambridge University Press. Morton FI (1983) Operational estimates of areal evapotranspiration and their significance to the science and practice of hydrology. Journal of Hydrology 66: 1–76. Oki T and Kanae S (2006) Global hydrological cycles and world water resources. Science 313: 1068--1072. Payne RE (1972) Albedo of the sea surface. Journal of the Atmospheric Sciences 29: 959--970. Peixoto JP and Oort AH (1996) The climatology of relative humidity in the atmosphere. Journal of Climate 9: 3443--3463. Penman HL (1948) Natural evaporation from open water, bare soil and grass. Proceedings of the Royal Society of London, Series A: Mathematical and Physical Sciences 193: 120--145. Peterson TC, Golubev VS, and Groisman PY (1995) Evaporation losing its strength. Nature 377: 687--688. Piao S, Friedlingstein P, Ciais P, De Noblet-Ducoudre N, Labat D, and Zaehle S (2007) Changes in climate and land use have a larger direct impact than rising CO2 on global river runoff trends. Proceedings of the National Academy of Sciences of the United States of America 104: 15242--15247. Porte´-Agel F, Parlange MB, Cahill AT, and Gruber A (2000) Mixture of time scales in evaporation: Desorption and self-similarity of energy fluxes. Agronomy Journal 92: 832--836. Priestley CHB and Taylor RJ (1972) On the assessment of surface heat flux and evaporation using large-scale parameters. Monthly Weather Review 100: 81--92. Raupach MR (2001) Combination theory and equilibrium evaporation. Quarterly Journal of the Royal Meteorological Society 127: 1149--1181. Rayner DP (2007) Wind run changes are the dominant factor affecting pan evaporation trends in Australia. Journal of Climate 20: 3379--3394. Roderick ML and Farquhar GD (2002) The cause of decreased pan evaporation over the past 50 years. Science 298: 1410--1411. Roderick ML, Rotstayn LD, Farquhar GD, and Hobbins MT (2007) On the attribution of changing pan evaporation. Geophysical Research Letters 34: L17403 (doi:10.1029/ 2007GL031166). Sene KJ, Gash JHC, and McNeil DD (1991) Evaporation from a tropical lake: Comparison of theory with direct measurements. Journal of Hydrology 127: 193--217. Seneviratne SI, Luethi D, Litschi M, and Schaer C (2006) Land–atmosphere coupling and climate change in Europe. Nature 443, doi:10.1038/nature05095. Shuttleworth WJ (1993) Evaporation. In: Maidment DR (ed.) Handbook of Hydrology, pp. 4.1--4.53. New York: McGraw-Hill. Shuttleworth WJ (2009) On the theory relating changes in area-average and pan evaporation. Quarterly Journal of the Royal Meteorological Society 135: 1230--1247. Shuttleworth WJ and Calder IR (1979) Has the Priestley–Taylor equation any relevance to forest evaporation? Journal of Applied Meteorology 18: 639--646. Stewart JB (1977) Evaporation from the wet canopy of a pine forest. Water Resources Research 13: 915–921. Sweers HE (1976) A nomogram to estimate the heat-exchange coefficient at the air– water interface as a function of wind speed and temperature: A critical survey of some literature. Journal of Hydrology 30: 375--401. Teuling AJ, Hirschi M, Ohmura A, et al. (2009) A regional perspective on trends in continental evaporation. Geophysical Research Letters 36(2): L02404. Thom AS and Oliver HR (1977) On Penman’s equation for estimating regional evaporation. Quarterly Journal of the Royal Meteorological Society 103: 345--357. Valentini R, Mateucci G, Dolman AJ, et al. (2000) Respiration as the main determinant of carbon balance in European forests. Nature 404: 861--865.

Evaporation in the Global Hydrological Cycle van der Molen MK, Dolman AJ, Waterloo MJ, and Bruijnzeel LA (2006) Climate is affected more by maritime than by continental land use change: A multiple scale analysis. Global and Planetary Change 54: 128--149. Wallace JS and Holwill CJ (1997) Soil evaporation from tiger-bush in south-west Niger. Journal of Hydrology 188–189: 426--442. Willett KM, Gillett NP, Jones PD, and Thorne PW (2007) Attribution of observed surface humidity changes to human influence. Nature 449, doi:10.1038/ nature06207.

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Wilson KB, Baldocchi DD, Aubinet M, et al. (2002) Energy partitioning between latent and sensible heat flux during the warm season at FLUXNET sites. Water Resources Research 38(12): 1294 (doi:10.1029/2001WR000989). Zhang L, Dawes WR, and Walker GR (2001) Response of mean annual evapotranspiration to vegetation changes at catchment scale. Water Resources Research 37: 701--708.

2.04 Interception AMJ Gerrits and HHG Savenije, Delft University of Technology, Delft, The Netherlands & 2011 Elsevier B.V. All rights reserved.

2.04.1 2.04.2 2.04.3 2.04.3.1 2.04.3.2 2.04.3.3 2.04.3.4 2.04.3.5 2.04.4 2.04.4.1 2.04.4.2 2.04.5 2.04.5.1 2.04.5.2 2.04.5.3 2.04.5.3.1 2.04.5.3.2 2.04.5.3.3 2.04.6 2.04.7 References

Introduction Importance of Interception Types of Interception Canopy Interception Forest Floor Interception Fog Interception Snow Interception Urban Interception Methods to Measure Interception Canopy Forest Floor Interception Models Conceptual Rutter Model Analytical Gash Model Stochastic Interception Models Poisson distribution Markov chains Gamma probability density function and transfer functions Consequences of Underestimating Interception for Hydrological Modeling and Water Resource Assessment Outlook

Nomenclature b Br c cp D Di E Ei Ei,c E li;c E ti;c Ei,f El Ep H I L m n nr,d nm p

constant in Rutter (1971) model (L1) Bowen ratio (–) canopy coverage (–) specific heat (L M T3 K1) drainage rate from the canopy (L T1) interception threshold (L T1) actual evaporation (L T1) interception evaporation (L T1) interception evaporation from canopy (L T1) evaporation from leaves (without trunk) (L T1) evaporation from the trunk (L T1) interception evaporation from forest floor (L T1) evaporation from lower basin (L T1) potential evaporation (L T1) sensible heat flux (M T3) interception process (L T1) number of elemental surface areas per unit ground (L2) mean number of raindrops striking an element (–) mean number of drops retained per element (–) number of rain days per month (–) days within a month ( ¼30.5) (–) throughfall coefficient (–)

pt P Pg P0g P00g q

r Sc Slc Stc Sf Si Sl Su Tf Ts t z b e k q h

89 90 90 90 91 91 91 92 93 93 94 94 94 95 95 95 96 98 98 99 99

trunk fraction coefficient (–) precipitation (L T1) gross precipitation (L T1) gross precipitation necessary for canopy saturation (L T1) gross precipitation necessary for trunk saturation (L T1) maximum amount of rain drops on element (–); specific humidity (M T1L1) amount of rain drops on element (–) storage of canopy (L) storage of leaves (without trunk) (L) storage of trunk (L) storage of forest floor (L) interception storage (L) storage of the lower basin (L) storage of the upper basin (L) throughfall (L T1) stemflow (L T1) mean volume of raindrops (L3) height (L) scaling factor (L T1) constant in Rutter (1971) model (–) latent heat of vaporization coefficient (L2 T2) density of water (M L3) potential temperature (K)

89

90

Interception

2.04.1 Introduction When it rains the entire surface becomes wet: trees, shrubs, grass, forest floor, footpaths, etc. Also in urban areas, roads and roofs become wet, sometimes forming pools of stagnant water. After rainfall has ceased these surfaces soon become dry again. This process is called ‘interception’. It is the part of the rainfall that is captured by surface storage (i.e., vegetation, roofs, etc.) before it can run off or infiltrate into the soil. The intercepted water generally evaporates during the event and shortly after the rainfall ceased, so that it can repeat its function during the next rainfall event. In the literature, interception is defined in different ways: sometimes as a stock, sometimes as a flux, or, more appropriately, as the entire interception process (Savenije, 2005). If only interception storage (Si [L]) is considered, interception is defined as the amount of rainfall which is temporarily stored on the Earth’s surface. Actually, this is the interception capacity or water-holding capacity. If interception is defined as a flux, then it is the intercepted water which evaporated over a certain time [L T1] during and after the event. When interception is considered as a process (I [L T1]), it is defined as the part of the rainfall flux which is intercepted on the wetted surface after which it is fed back to the atmosphere. The interception process equals the sum of the change of interception storage (Si) and the evaporation from this stock (Ei):

I¼

dSi þ Ei dt

ð1Þ

The timescale of the interception process is in the order of 1 day. After 1 day, it is fair to assume for most climates that the first term on the right-hand side in Equation (1) approaches zero, and I ¼ Ei. Of course, in the case of snow under cold climates, this may take longer. How much of the precipitation is intercepted depends on several factors, which can be divided into three groups:

•

Vegetation characteristics. Large vegetation types, such as trees, have a high aerodynamic roughness, causing high potential evaporation rates. Grasses, crops, or bushes, on the other hand, have a much lower roughness and thus do not have as high potential evaporation rates. The storage capacity also depends on the vegetation type. The shape of the leaves, the thickness, the density (leaf area index), and the configuration of the branches determine how much water can be stored. For example, the capacity of a coniferous or a deciduous tree is different (e.g., Rutter et al., 1975; Baird and Wilby, 1999; Bryant et al., 2005; Toba and Ohta, 2005). Although intuitively one might think that a deciduous tree can hold more water in its bucket-like leaves, a coniferous tree can hold much more water by adhesion. Furthermore, it is also important to take the seasonality into account. Deciduous trees lose their leaves in the dormant season, causing a large reduction in the canopy storage capacity. Vegetation also determines the amount of understorey growth and forest floor. The forest floor of different vegetation types can have significantly

•

•

different interception behavior (e.g., a thick needle layer or a thin leaf litter layer). Rainfall characteristics. Rainfall has a large influence on the interception process. The rainfall frequency is a major determining factor. It makes a big difference if rainfall occurs as one continuous storm or as a sequence of several small events with dry spells in between. Even if the total rainfall depth is the same, the last scenario intercepts much more rainwater, because between the events the storage can be (partly) emptied by evaporation and thus more storage is available. Second, the rainfall intensity is important, although there is no consensus in literature. Horton (1919) and Wang et al. (2007) concluded that the interception capacity is lower at higher intensity because high rainfall intensities cause splashing and shaking of leaves. On the other hand, Aston (1979) and Keim et al. (2006b) noted the opposite: high rainfall intensities coincide with high storage capacities, due to dynamic storage. Evaporative demand. If the potential evaporation (i.e., open water evaporation) is high, the intercepted water can evaporate more easily during and after the event. Wind plays an important role in removing moisture from the surface providing a higher vapor deficit, particularly in the canopy. Moreover, the roughness of the vegetation increases the evaporative power, by causing turbulence which makes it easier to take up the intercepted water. However, wind can also reduce the amount of interception by reducing the storage. Horton (1919), Klaassen et al. (1996), and Ho¨rmann et al. (1996) noted that with increasing wind speed the measured storage capacity is less, due to the fact that the wind shakes the rainwater off the leaves.

Of the above three factors, the rainfall characteristics are most dominant for evaporation from interception. Although both the storage capacity (mainly vegetation characteristic) and the available energy form a constraint to the evaporation flux per event, the number of events is a more important factor. This is confirmed by the sensitivity analysis of Gerrits et al. (2009c).

2.04.2 Importance of Interception Although most surfaces can store only a few millimeters of rainfall, which is often not much in comparison to other stocks in the water balance, interception is generally a significant process. The impact becomes evident at longer timescales. Although interception storage is generally small, the number of times that the storage is filled and depleted can be so large that the interception flux is generally of the same order of magnitude as the transpiration flux. In addition, the interception process smooths the rain intensities, causing more gradual infiltration. Interception redistributes the rainfall as well. Some parts of a field receive less water due to interception, whereas other parts receive more due to funneling of the vegetation (e.g., Germer et al., 2006; Gerrits et al., 2009b). Subsequently, this has an influence on the soil moisture patterns, and this is again important for flood generation (Roberts and Klingeman, 1970). Besides the hydrological effects, there are influences on the nutrient cycle of a forest, and on agricultural applications.

Interception

For example, interception affects the efficiency of insecticides and fertilizers (Aston, 1979). Besides, fire retardants are more effective if they are stored by vegetation. Finally, interception may reduce soil erosion by preventing rain drops to directly hit and erode the soil layer (Walsh and Voigt, 1977), although in the case of canopy interception the opposite can be true due to the formation of larger rain drops with a higher impact on the forest floor.

2.04.3 Types of Interception As already stated in Section 2.04.1, it is possible to define an infinite number of interception types. In principle, every surface that can store water can be considered as an interception type. In this chapter, we focus on the major types, mainly occurring in a natural environment plus some special mechanisms. However, more often than not, it is a combination of mechanisms. For example, in a forest, it is likely that a part of the rainfall is intercepted by the canopy of a tree, while the remaining part can be intercepted by epiphytes on the branches and/or bark, and, finally, the understorey and forest floor intercept the throughfall before infiltration starts.

2.04.3.1 Canopy Interception Canopy interception is the rainwater that is stored on the leaves and branches of a tree which is subsequently evaporated. This interception can be calculated by measuring rainfall above the trees or measured in an open area nearby (gross

Precipitation

Canopy interception

91

rainfall Pg) and subtracting the throughfall (Tf) and stemflow (Ts) (Figure 1):

Ei;c þ

dSc ¼ Pg  Tf  Ts dt

ð2Þ

Many research studies have been carried out on canopy interception. In Table 1 an overview is given. We can see in the table and also in tables in Kittredge (1948), Zinke (1967), and Breuer et al. (2003) that there is a large difference in the canopy interception by deciduous and coniferous trees (e.g., Kittredge, 1948; Bryant et al., 2005; Toba and Ohta, 2005). Not only because deciduous trees lose their leaves, but also because the leaf area of coniferous trees is much larger than of deciduous trees; coniferous trees can store much more water. Furthermore, leaves may swing over when they become too heavy, causing a (sudden) decrease of the storage capacity. However, Herbst et al. (2008) found counterintuitive results, where higher evaporation rates were found in deciduous trees in winter caused by rougher aerodynamics of the bare canopy and deeper penetration of the wind. In most cases, the storage of water on the branches is small; however, in some environments, the branches can be overgrown by epiphytes. Pypker et al. (2006) showed that in a Douglas fir forest the canopy water storage can potentially be increased by 41.3 mm and Ho¨lscher et al. (2004) found that epiphytes can account for 50% of the storage capacity. However, this large increase in storage capacity is not necessarily resulting in high interception values (storage þ evaporation), because the water uptake and release by the epiphytes is delayed. It takes a while to saturate the epiphytes, and already before saturation, runoff generation can take place. Successively, after wetting, the drying of the epiphytes takes much longer than drying of the canopy, causing less storage to be available. Another special type of canopy interception is interception by agricultural crops. In essence, there is no difference between crops and other vegetation types. They both can store water up to a certain threshold and then drain water to the floor as throughfall. However, whereas vegetation has a gradual seasonal pattern (summer vs. winter), crops have a phenological growth cycle (seeding to harvesting) which is therefore more abrupt. Hence, when modeling crop interception the appropriate description of the variation in the storage capacity is important.

2.04.3.2 Forest Floor Interception

Throughfall Stemflow Forest floor interception Infiltration Figure 1 Two major interception types in the natural environment.

Forest floor interception is the part of the throughfall that is temporarily stored in the top layer of the forest floor and successively evaporated within a few hours or days during and after the rainfall event. The forest floor can consist of short vegetation (like grasses, mosses, bushes, and creeping vegetation), litter as described by Hoover and Lunt (1952) as the litter and fermentation (L and F) layer (i.e., leaves, twigs, and small branches), or bare soil. Although the latter seems to have an overlap with soil evaporation, we distinguish them by the fact that soil evaporation refers to the water that is stored in the root zone (Groen and Savenije, 2006).

92

Interception

Table 1 Canopy interception values in literature, with Sc,max the water storage capacity and Ei,c the interception evaporation as percentage of gross precipitation Source

Specie

Location

Sc,max (mm)

Rutter et al. (1975)

Corsian pine (Pinus nigra) Douglas fir (Pseudotsuga menziesii) Norway spruce (Picea abies) Hornbeam (Carpinus betulus) Oak (Quercus robur)

United United United United United

Gash and Morton (1978) Gash et al. (1980)

Scots pine (Pinus sylvestris) Sitka spruce (Picea sitchensis) Scots pine (Pinus sylvestris) Beech (Nothofagus) Acacia auriculiformis

United Kingdom United Kingdom United Kingdom New Zealand Indonesia

1.05 1.2 1.5 1.0 (leafy) 0.65 (leafless) 0.875 (leafy) 0.275 (leafless) 0.8 0.75–1.2 1.02 1.5 (leafy) 1.2 (leafless) 0.5–0.6

Norway spruce (Picea abies) Beech (Asperulo-fagetum) Pinus pinaster Eucalyptus globulus Tamaulipan thornscrub Loblolly (Pinus taeda) & shortleaf pine (Pinus echinata) Longleaf pine (Pinus palustris) Scrub oak (Quercus berberidifolia) White oak (Quercus alba) & shortleaf pine (Pinus echinata) & loblolly pine (Pinus palustris) Hardwood Larc (Larix cajanderi) Red pine (Pinus sylvester) Red pine (Pinus densiflora) Sawtooth oak (Quercus acutissima) Oak (Quercus serrata) Rain forest

France Germany Portugal Portugal Mexico USA (GA)

Rowe (1983) Bruijnzeel and Wiersum (1987) Viville et al. (1993) Ho¨rmann et al. (1996) Valente et al. (1997) Navar et al. (1999) Bryant et al. (2005)

Toba and Ohta (2005)

Cuartas et al. (2007)

Kingdom Kingdom Kingdom Kingdom Kingdom

Ei,c (%) 35 39 48 36 18

27–32 42 35 (leafy), 22 (leafless) 11–18

1.97

34.2 18 10.8 17.1 18.9 22.3

USA (GA) USA (GA) USA (GA)

1.70 1.40 1.58

17.6 17.4 18.6

USA (GA) Siberia Siberia Japan Japan Japan Brazil

0.98

17.7 29 36 13–17 24 18 13–22

1.28 (leafy) 0.84 (leafless) 0.41 0.21

1.0

See also tables in Kittredge (1948), Zinke (1967), and Breuer et al. (2003).

In Table 2 some results are presented of previous work on forest floor interception.

moisture. These instruments suffer from various limitations. An overview of fog collectors can be found in Bruijnzeel et al. (2005).

2.04.3.3 Fog Interception A special type of interception is fog interception or cloud interception. Vegetation can intercept not only rain, but also moisture (in the form of small water droplets) from the air. Fog can occur due to different processes. Bruijnzeel et al. (2005) distinguished nine types: radiation fog, sea fog, stream fog, advection fog, ice fog, coastal fog, valley fog, urban fog, and mountain fog. Fog interception is mainly important in tropical montane environments (table in Bruijnzeel (2005): 6–53% of rainfall), and can also play a significant role in semi-arid regions near the coast (e.g., Hursh and Pereira, 1953; Hutley et al., 1997; Hildebrandt et al., 2007). In both environments, the main problem with fog interception studies is to measure precipitation and throughfall (Equation (2)), which is especially important because fog deposition can be twice as high as normal rainfall. Since conventional rain gauges are not suitable to measure fog deposition, special fog collectors have been developed with often wire meshes to intercept the

2.04.3.4 Snow Interception Snowfall is also intercepted by trees. Especially, coniferous trees can store so much snow, that they collapse under its weight. As an example, Storck et al. (2002) found in a Douglas-fir-dominated forest that up to 60% of the snowfall was intercepted, equaling 40 mm of snow water equivalent (swe). The storage of snow on the canopy is different from rain. For rainfall interception the storage capacity is mainly a function of the leaf surface area, whereas for snow interception the branch strength and canopy shape are more important (Ward and Trimble, 2004). Furthermore, the snow storage is also dependent on the temperature. If snow falls with temperatures close to freezing point, the cohesion of snow is higher causing more snow to be accumulated on the canopy (Ward and Trimble, 2004). Another difference between rainfall interception and snow interception is the way in which interception storage is depleted. Rainfall interception is a real threshold process,

Interception

93

Table 2 Forest floor interception values in literature, with the water storage capacity Sf,max and the interception evaporation Ei,f as percentage of net precipitation (i.e., throughfall) Source

Forest floor type

Location

Haynes (1940) Kittredge (1948) Beard (1956) Helvey (1964) Brechtel (1969)

Kentucky bluegrass (Poa pratensis) Californian grass (Avena, Stipa, Lolium, Bromus) Themeda and Cymbopogon Poplar Scot’s pine Norway spruce Beech Oak Shorea robusta and Mallotus philippensis Pinus roxburghii and Quercus glauca Pinus roxburghii Quercus leucotrichophora and Pinus roxburghii Quercus floribunda and Quercus leucotrichophora Quercus lanuginosa and Quercus floribunda Blue stem Andropogon gerardi Vitman Pine (Pinus sylvestris) Beech (Fagus sylvaticus) Bracken litter (Pteridium aquiliunum) Norway spruce Sitka spruce Beech (Asperulo-Fagetum) Pinus radiata Eucalyptus Douglas fir Peble mulch (5–9 cm) Peble mulch (2–6 cm) Cryptomeria japonica Lithocarpus edulis Grass (Aristida divaricata) Woodchips (Pinus) Poplar leaves (Populus nigra)

? USA (CA) South Africa USA (NC) USA (NY) USA (NY) USA (NY) USA (NY) India India India India India India USA (TX) United Kingdom United Kingdom United Kingdom Scotland Scotland Germany Australia Australia Netherlands China China Japan Japan Mexico Mexico Mexico

Pathak et al. (1985)

Clark (1940) in Thurow et al. (1987) Walsh and Voigt (1977) Pitman (1989) Miller et al. (1990) Thamm and Widmoser (1995) Putuhena and Cordery (1996) Schaap and Bouten (1997) Li et al. (2000) Sato et al. (2004) Guevara-Escobar et al. (2007)

a

Sf,max (mm)

Ei,f (%) 56a 26a 13a 34 21 16 16 11 11.8 7.8 9.6 10.6 11.0 11.3 57–84

0.6–1.7 0.9–2.8 1.67

2.5–3.0 2.78 1.70 0.281 0.526 0.27–1.72 0.67–3.05 2.5 8 2.3

18a 16a 12–28

0.23 mm d1 11.5a 17.4a

% of gross precipitation instead of net precipitation.

whereby throughfall starts when the storage capacity is exceeded. The storage capacity is then emptied by evaporation. Snow, on the other hand, can only be removed from the canopy by three ways: sublimation, mechanical removal (sliding leading to mass release), and melt water drip (Miller, 1966).

Island is mainly caused by the (relatively warm) buildings that block the cold night sky. Furthermore, the thermal properties of a city are different: concrete and asphalt have much higher heat capacities than forests and also the surface radiative properties differ (e.g., albedo and emissivity). The lack of vegetation in urban areas, which reduces cooling by transpiration, also causes a difference in the energy balance.

2.04.3.5 Urban Interception Most hydrological studies focus on natural environments and not on urbanized areas, which is also the case for interception studies. However, recently, with the increasing interest for alternative sources of water for nonpotable domestic use (socalled ‘gray water’), water balance studies on (interception) evaporation in urban areas increased (Grimmond and Oke, 1991; Ragab et al., 2003; Gash et al., 2008; Nakayoshi et al., 2009). The difference between urban and rural interception is not only that the typical storage capacities of buildings, roads, etc., are unknown, but also that the entire energy balance is different in a city. Oke (1982) discovered the so-called ‘Urban Heat Island’, that is, higher temperatures in urban areas compared to the surrounding rural areas. The Urban Heat

2.04.4 Methods to Measure Interception 2.04.4.1 Canopy There exist already many methods to measure canopy interception. The most-often used method is by measuring rainfall above the canopy and subtract throughfall and stemflow (e.g., Helvey and Patric, 1965). However, the problem with this method is that the canopy is not homogeneous, which causes it to be difficult to obtain representative throughfall data. Using multiple rain gauges under the canopy (Helvey and Patric, 1965; Keim et al., 2005; Gerrits et al., 2009b) reduces this problem. Sometimes the collectors are moved to achieve a better representation of throughfall (e.g., Lloyd and Marques, 1988; Tobo´n-Marin et al., 2000; Manfroi et al., 2006; Ziegler

94

Interception

et al., 2009). Another method to avoid the problem with the spatial distribution of the canopy was introduced by Calder and Rosier (1976) and applied by, for example, Shuttleworth et al. (1984), Calder et al. (1986), and Calder (1990). They covered the forest floor with plastic sheets and collected the throughfall. The disadvantage of this method is that for long periods irrigation is required, because otherwise, in the end, the trees will dry out and may even die due to water shortage. The method by Hancock and Crowther (1979) avoided these problems, by making use of the cantilever effect of branches. If leaves on a branch hold water, it becomes more heavy and will bend. By measuring the displacement, it is possible to determine the amount of intercepted water. Huang et al. (2005) refined this method by making use of strain gauges. However, the disadvantages of these methods are that only information about one single branch is obtained and it is quite laborious to measure an entire tree. Edwards (1986), Fritschen and Kinerson (1973), and Storck et al. (2002) made use of weighing lysimeters with trees. Although interception of a whole tree is measured with this method, the big disadvantage of this method is that it is expensive and destructive. Friesen et al. (2008) developed a nondestructive method to measure canopy interception of a whole tree. With mechanical displacement sensors, Friesen et al. (2008) measured the stem compression due to interception water, which is an integration of the whole canopy. However, although this method looks promising, it is still under development. A totally different way of measuring canopy interception of a forest plot is to make use of ray attenuation. Calder and Wright (1986) used the attenuation of gamma rays. They transmitted from a tower gamma-rays through the canopy at different heights and measured the gamma-ray density at a receiving tower. The ratio between transmitted and received gamma-ray density during dry conditions is successively compared to this ratio during a rainfall event. This gives an estimate of the amount of water stored on the canopy over time. Although the method gives interception estimated of an entire forest, the method becomes inaccurate under windy conditions. Furthermore, safety standards inhibits unattended use of this method. Bouten et al. (1991) overcame this problem by making use of microwave attenuation. It appears to be a suitable method to measure canopy wetness, although it is an expensive method. Evaporation can also be measured by flux measurements. By measuring temperature (y) and specific humidity (q) at several heights (z) above the canopy, one can calculate the Bowen ratio (Br), which is the sensible heat flux, H, divided by the latent heat flux (lE):

Br ¼

cp dy=dz H ¼ rlE ldq=dz

ð3Þ

Combined with the energy balance, evaporation can be calculated (Gash and Stewart, 1975). The main difficulty with the Bowen ratio method is to measure the humidity gradient more accurately (Stewart, 1977). Another method is the eddy covariance technique, where the net upward or downward flux is determined by fast-response three-dimentional (3D) wind speed measurements combined with a concentration

measurement. This concentration can be humidity, temperature, or CO2 concentrations (Amiro, 2009).

2.04.4.2 Forest Floor In the literature, little can be found on forest floor interception, although some researchers have tried to quantify the interception amounts. Generally, these methods can be divided into two categories (Helvey and Patric, 1965): 1. lab methods, whereby field samples are taken to the lab and successively the wetting and drying curves are determined by measuring the moisture content and 2. field methods, whereby the forest floor is captured into trays or where sheets are placed underneath the forest floor. An example of the first category is that of Helvey (1964), who performed a drainage experiment on the forest floor after it was saturated. During drainage, the samples were covered, and after drainage had stopped (24 h), the samples were taken to the lab, where the samples were weighed and successively dried until a constant weight was reached. By knowing the oven dry weight of the litter per unit area and the drying curve, the evaporation from interception could be calculated. In this way, they found that about 3% of the annual rainfall evaporated from the litter. Similar work was done by Bernard (1963), Walsh and Voigt (1977), and Sato et al. (2004). However, what they all measured was not the flux, but the storage capacity. Another example of lab experiments was carried out by Putuhena and Cordery (1996). First, field measurements were carried out to determine the spatial variation of the different forest floor types. Second, storage capacities of the different forest floor types were measured in the lab using a rainfall simulator. Finally, the lab experiments were extrapolated to the mapping step. In this way, Putuhena and Cordery (1996) found average storage capacities of 2.8 mm for pine and 1.7 mm for eucalyptus forest floors. Moreover, GuevaraEscobar et al. (2007) made use of a rainfall simulator. Examples of the second category have been, for example, carried out by Pathak et al. (1985), who measured the weight of a sample tray before and after a rainfall event. They found litter interception values of 8–12% of the net precipitation. In addition, here, they measured the storage capacity, rather than the flux. Schaap and Bouten (1997) measured the interception flux by the use of a lysimeter and found that 0.23 mm d1 evaporated from a dense Douglas fir stand in early spring and summer. Also, Brechtel (1969) and Thamm and Widmoser (1995) made use of lysimeters. Brechtel (1969) manually measured the infiltrated water and Thamm and Widmoser (1995) developed an automatic and more sophisticated method, whereby the suction under the forest floor is controlled by a tensiometer. Gerrits et al. (2007) developed a method whereby both the forest floor interception and the infiltrated water are continuously weighed in suspended trays with strain gauges. In Figure 2 the schematic setup is shown. Measurements with sheets were done, for example, by Li et al. (2000), who found that pebble mulch intercepts 17% of the gross precipitation. Miller et al. (1990) found comparable results (16–18%) for a mature coniferous plantation in Scotland.

Interception

95

three parts:

Eint

Precipitaion

1. free throughfall, that is, throughfall, which did not touch the canopy at all (pPg), 2. trunk input (ptPg), and 3. canopy input ((1  p  pt)Pg). Litter Su Geotextile El

Weighing device

Infiltration

The rain that falls on the canopy can drain to the ground (i.e., canopy drainage, D), or evaporate ðEli;c Þ, or it can be stored on the canopy ðSlc Þ:

ð1  p  pt Þ

Z

Pg dt ¼

Z

Ddt þ

Z

Eli;c dt þ

Z

dSlc

ð4Þ

Sl

The rain that falls on the trunk can evaporate from the trunk ðEti;c Þ, or drain in the form of stemflow (Ts), or it can be stored on the trunk ðStc Þ:

Valve

Z Figure 2 Forest floor interception device by Gerrits AMJ, Savenije HHG, Hoffmann L, and Pfister L (2007) New technique to measure forest floor interception – an application in a beech forest in Luxembourg. Hydrology and Earth System Sciences 11: 695–701.

2.04.5 Interception Models In literature, several models have been developed to simulate forest interception. Almost all of these models are concentrated on canopy interception, sometimes including stem interception (Table 3). In principle, these models can be expanded to include forest floor or any surface interception as well. The most often used interception models are the conceptual model of Rutter et al. (1971) (Section 2.04.5.1) and the analytical model of Gash (1979) (Section 2.04.5.2) or revisions of these models. Furthermore, there exist some stochastic models, which will be described in Section 2.04.5.3. In Table 3 an overview and summary of the models are given. A more detailed overview and comparison can be found in Muzylo et al. (2009).

2.04.5.1 Conceptual Rutter Model The conceptual framework of the original Rutter model is depicted in Figure 3. As can be seen the rainfall is divided into Table 3

Characteristics of interception models

Main author

Model type

Rutter Gash C alder De Groen

Conceptual Analytical Stochastic Concept./ stoch. Concept./ stoch.

Keim

Interception element: canopy

stem

x x x x

x x

x

x

Timescale

forest floor

x

rhourly event rhourly monthly 6-hourly

pt

Pg dt ¼

Z

Ts dt þ

Z

Eti;c dt þ

Z

dStc

ð5Þ

with Ei;c ¼ Eli;c þ Eti;c and Sc ¼ Slc þ Stc for the total canopy interception. The evaporation from the wet canopy is calculated with the Penman equation (Penman, 1948). Because the canopy is not always completely wet ðSlc o Slc;max Þ, the actual evaporation rate can be calculated by the fraction of the potential evaporation: Ep Slc =Slc;max . The same concept is applied for the trunks. However, for the determination of the potential evaporation of the trunks, the potential evaporation of the canopy is multiplied with and extra constant E. Stemflow is modeled as a threshold process, whereby no stemflow is generated when Stc o Stc;max , and when the threshold is exceeded stemflow equals the difference between Stc and Stc;max . Canopy drainage is modeled in a similar way; however, when the threshold Slc;max is exceeded, drainage is defined as

h i D ¼ Ds exp bðSlc  Slc;max Þ

ð6Þ

with Ds being the rate of drainage when the canopy is saturated and b [L1] an empirical coefficient. Valente et al. (1997) revised the original Rutter model, to model interception in a more realistic way for sparse canopies. The main drawbacks of the original model were the partitioning of free throughfall and canopy input, and the conceptual error that evaporation from interception can theoretically be higher than potential evaporation (Valente et al., 1997). Therefore, they divided the conceptual model into two areas: a covered area (c) and an uncovered area (1  c). Second, in the revised Rutter model, only water can reach the trunk after it has flowed through the canopy as a part of the canopy drainage. Water which is not drained by the trunk is directly dripping to the ground. The final change was made that evaporation from the saturated canopy is not equal to the potential evaporation, but is reduced by a factor 1  E (0oEo1). The remaining energy ((E)Ep) is then available for evaporating water from the saturated trunk (Figure 4).

96

Interception

Canopy evaporation S lc ⎧ ,S lc< S lc,max E ⎪ E li,c = ⎨ p S lc,max ⎪ ,S lc ≥ S lc,max Ep ⎩

Canopy input ( 1−p−pt )Pg

Scl

Gross rainfall Pg

Free throughfall p Pg

Trunk evaporation S tc ⎧ ,S tc < S tc,max ⎪εEp t t E i,c = ⎨ S c,max ⎪ ,S tc ≥ S tc,max ⎩ εEp

Trunk input pt Pg

Sct

Scl, max

Sct,max

Drainage D = Ds exp[b (Scl- Scl,max]

Throughfall, Tf

Stemflow, Ts

Figure 3 Conceptual framework of the Rutter model. Modified from Valente F, David JS, and Gash JHC (1997) Modelling interception loss for two sparse eucalypt and pine forest in central Portugal using reformulated Rutter and Gash analytical models. Journal of Hydrology 190: 141–162.

2.04.5.2 Analytical Gash Model The original Gash model is conceptually the same as the Rutter model (see Section 2.04.5.1); however, it does not require meteorological data of high temporal resolution (hourly) and requires less computation time. The main assumption of the Gash model is that it is possible to represent the real rainfall pattern by different discrete rainfall events, each consisting of three phases: 1. wetting phase, 2. saturation phase, and 3. drying phase (long enough to dry the entire canopy). Similar to the Rutter model, rainfall is divided into canopy input (1  p  pt), free throughfall (p), and trunk input (pt). The Gash model makes a distinction between storms which are not large enough to saturate the canopy ðPg o P0g : m stormsÞ and storms which are large enough to saturate the canopy ðPg  P0g : n stormsÞ. The amount of gross rainfall necessary to saturate the canopy is P0g (see Table 4). Interception evaporation is then calculated for the canopy and the trunk. Although the original Gash model appears to work fine for several types of forests, it contains some weaknesses for modeling sparse forests, similar to the Rutter model. Hence, Gash et al. (1995) revised their existing model according to the revised Rutter model (Rutter et al., 1975). An overview of the formulas of the revised Gash model can be found in Table 4.

2.04.5.3 Stochastic Interception Models

areas, which all have the same probability to be struck by raindrops. The Poisson probability of an element to be struck by r drops equals

Pr ¼

m r m e r!

ð7Þ

with m the mean number of raindrops striking an element per storm. If an element can hold q raindrops, the mean number of drops per element (n) can be expressed as

n¼

q X

r  Pr þ q  Pðr 4 qÞ

ð8Þ

r¼0

¼qþ

q X

Pr  ðr  qÞ

ð9Þ

r¼0

with P (r4q) the probability of elements being struck by more P than q drops and is equal to 1  qr¼ o Pr. To upscale from elemental area to canopy area, the number of elemental surface areas per unit ground (L) is required and the mean volume of raindrops (u):

Sc ¼ nuL

ð10Þ

Sc;max ¼ quL

ð11Þ

Pg ¼ muL

ð12Þ

Evaporation is then obtained by (with dScdEi,c ¼  1)

2.04.5.3.1 Poisson distribution Calder (1986) developed a stochastic interception model, where he assumes that a tree consists of several elemental

dn dSc dn 1 ¼  ¼ dEi;c dEi;c dSc uL

ð13Þ

Interception Gross rainfall Pg

97

Interception from evaporation: Ei,c = c·( E li,c + E ti,c )

Uncovered area input Pg

Uncovered area (1−c)

Covered area input Pg

Covered area c

S lc

⎧ ⎪ (1−ε)Ep l S lc,max E i,c = ⎨ ⎪ (1 − ε) Ep ⎩

Free throughfall Pg

Trunk evaporation:

Canopy evaporation: ,S lc < S lc,max ,S lc ≥ S lc,max

S tc ⎧ ⎪ εEp t S c,max i,c = ⎨ ⎪ εEp ⎩

Et

,S tc,< S tc,max ,S lc, ≥ S tc,max

Scl Scl, max

Drainage Dc = d(S lc−S lc,max)/dt

Drip D i,c =( 1−pd ) Dc

Trunk input pd Dd

Sct Sct, max

Trunk drainage Dt,c = d(Sct- Sct, max)/dt

Throughfall, Tf (1−c)Pg+c Di,c

Stemflow, Ts c Dt,c

Figure 4 Conceptual framework of the revised Rutter model. Modified from Valente F, David JS, and Gash JHC (1997) Modelling interception loss for two sparse eucalypt and pine forest in central Portugal using reformulated Rutter and Gash analytical models. Journal of Hydrology 190: 141–162.

The Calder model is very simple and describes the threshold behavior of interception very well; however, it is difficult to upscale from drop size scale to forest size scale. This hinders the applicability of the model.

2.04.5.3.2 Markov chains Groen and Savenije (2006) developed a monthly interception model based on a daily interception model and the daily rainfall characteristics. They assumed interception on a daily

98 Table 4

Interception Components of interception of the original Gash (1979) model and the revised Gash et al. (1995) model for sparse canopies Original Gash (1979)

Amount of gross rainfall necessary to saturate the canopy ðP 0g Þ and trunk ðP00g Þ

Revised (sparse canopy) Gash et al. (1995)

"

P 0g

 g Sc;max p P E ¼  p ln 1  ð1  p  pt ÞP g E

P 00g ¼ Stc;max =pt Evaporation from canopy interception ðE li;c Þ: 1. for m storms ðPg o P 0g Þ 2. for n storms ðP g  P 0g Þ Evaporation from trunk interception ðE ti;c Þ: 1. for q storms ðP g  P 00g Þ 2. for m þ n  q storms ðPg o P 00g Þ

ð1  p  pt Þ

# P 0g

" # g p P Sc;max ð1  EÞE ¼ g  p c ln 1  P ð1  EÞE

g Stc;max P 0 P 00g ¼   p pt c þ P g P g  ð1  EÞE

Pm

Pm P g;j " j¼1 #  p Pn ð1  EÞE 0 0 c nP g þ j¼1 ðP g;j  P g Þ g P

Pg;j  p Pn E 0 ðP g;j  P 0g Þ nð1  p  pt ÞP g þ  P g j¼1

c

qStc P Pg;j pt mþnq j¼1

qStc "

j¼1

pt c 1 

#  p Pn ð1  EÞE 0 ðP  P Þ g g;j j¼1 g P

Modified from Valente F, David JS, and Gash JHC (1997). Modelling interception loss for two sparse eucalypt and pine forest in central Portugal using reformulated Rutter and Gash analytical models. Journal of Hydrology 190:141–162.

scale as (Savenije, 1997, 2004)

Markov probabilities to model monthly interception based on daily information.

Ei;d ¼ minðDi;d ; Pg;d Þ

ð14Þ

The probability distribution of rainfall on a rain day can be described as

  Pg;d 1 f i;d ðPg;d Þ ¼ exp  b b

ð15Þ

with b being the scaling factor, equal to the expected rainfall on a rain day, which can be expressed as

b¼

Pg;m Eðnr;d =nm Þ

ð16Þ

with Pg,m being the monthly rainfall and nr,d and nm the number of rain days per month and amount of days per month, respectively. The number of rain days per month can be expressed by the use of Markov properties. Being p01 the Markov properties of the transition from a dry day to a rain day, and p11 the probability of a rain day after a rain day:

nr;d ¼ nm

p01 1  p11 þ p01

ð17Þ

Multiplying Equations (14) and (15) and successively integrating results in monthly evaporation from interception:

Ei;m ¼ Eðnr;d jnm Þ

ZN

Ei;d  f i;d ðPd Þd Pd

ð18Þ

2.04.5.3.3 Gamma probability density function and transfer functions Keim et al. (2004) developed a stochastic model to obtain from 6-hourly rainfall to 6-hourly throughfall for extreme events. They made use of the gamma probability density function (PDF; for 6 hourly rainfall):

Pga1 ePg =y Tf  100% ¼ Pg GðaÞya

ð20Þ

The parameters a and y can be estimated by dividing the 6-hourly rainfall in ranges and find the best-fit sets. After downscaling the rainfall and throughfall data, rainfall is transferred through the canopy by a linear system convolution to obtain high-resolution throughfall data, which allows one to investigate the effect of intensity smoothing:

Tf ðtÞ ¼

Zt

PðtÞgðt  tÞdt

ð21Þ

0

with the transfer function g(t  t). Keim et al. (2004) found that the transfer function can be best described with the exponential distribution:

0

   Di;d ¼ Pm 1  exp b

ð19Þ

Hence, the model of Groen and Savenije (2006) is a parsimonious model with only one measurable parameter, and

gðtÞ ¼ a e at

ð22Þ

By coupling the stochastic model with the the intensity smoothing transfer function, effects of forest canopies on extreme rainfall events can be investigated.

Interception

2.04.6 Consequences of Underestimating Interception for Hydrological Modeling and Water Resource Assessment Hydrologists often consider precipitation as the start of the hydrological cycle. After a rainfall event, the first separation point in the cycle is on the Earth surface. Part of the rainwater is intercepted by the vegetation or ground surface and the remainder infiltrates into the unsaturated zone or runs off. The part of the rainfall that is intercepted successively evaporates from the temporary storage. This first separation point in the hydrological cycle is not always considered a significant process. This is partly due to the technical difficulties that are inherent to interception measurements (Lundberg et al., 1997; Llorens and Gallart, 2000), but it is also generally considered a minor flux, although previous studies tell us that interception can amount to 10–50% of the precipitation depending on the vegetation type (Klaassen et al., 1998). Even then, these studies mostly refer to canopy interception only. If forest floor interception is taken into account as well, the percentage is substantially higher. Furthermore, it is often stated that interception is particularly not important for the generation of floods. This is not true. Interception strongly influences the antecedent soil moisture conditions, which are very important for the generation of floods (Roberts and Klingeman, 1970). Still, interception is regularly (partly) disregarded in hydrological models, or taken as a fixed percentage of the precipitation. As a result, after model calibration, interception is generally compensated by other processes such as transpiration, soil evaporation, or even recharge (Savenije, 2004). Zhang and Savenije (2005) showed that the hydrograph at the outlet of the Geer basin in Belgium improved significantly when interception was included in a rainfall–runoff model using the representative elementary watershed (REW) approach. Both the Nash–Sutcliffe efficiency and the percentage bias improved. They also showed that, in calibration, the soil moisture storage capacity compensated for the neglect of the interception process. Keim et al. (2006a) investigated the effects of (canopy) interception. They looked at the influence on the subsurface stormflow generation and concluded that interception caused a delay in the onset of subsurface stormflow, lowered and delayed stormflow peaks, and decreased total flow and the runoff ratio. They also found that simply reducing the rainfall by a constant factor did not result in a satisfactory peak flow response. Fenicia et al. (2008) looked at the change in the movement of the Pareto front when stepwise new processes were included in a variable model structure. They concluded that when interception was included and especially, when spatially distributed interception was included, the Pareto front moved significantly to the origin. Hence, their conclusion was that interception is an important process and should therefore be included in hydrological models.

2.04.7 Outlook More than 2000 articles have been published on interception studies (source: Scopus and ISI Web of KnowledgeSM) and still

99

new articles are being published. Most of these articles focus on canopy interception and describe in detail the process for different tree species in different climates, resulting in long reference tables as, for example, presented in Table 1 and by Breuer et al. (2003). Although this information is of high value for modeling purposes, it would have been more logical if these tables had also been available for the other types of interception, such as described in Section 2.04.3. Especially, since generally more than one mechanism occur, these mechanisms interact (see, e.g., between canopy and forest floor interception (Gerrits et al., 2009a). It would really be a way forward, if a broader scope was systematically considered in interception studies. Although a more balanced database on interception values will help, it is not the complete solution for hydrological modeling. Often, experimental results are site and time specific. Therefore, it is difficult to upscale literature values on interception for catchment modeling. This problem may be solved by considering the energy balance. If we would know how the available energy is partitioned over the different fluxes and compartments, we would be able to determine interception evaporation as well. However, this would require intensive field experiments where both the energy fluxes and the evaporation processes are measured simultaneously. Remote sensing could provide the necessary spatial and temporal information on energy partitioning. Through a combination of methods, interception could be more adequately incorporated in hydrological models.

References Amiro B (2009) Measuring boreal forest evapotranspiration using the energy balance residual. Journal of Hydrology 366(1–4): 112--118. Aston A (1979) Rainfall interception by eight small trees. Journal of Hydrology 42 (3–4): 383--396. Baird AJ and Wilby RL (eds.) (1999) Eco-Hydrology-Plants and Water in Terrestrial and Aquatic Environments. London: Routledge. Beard JS (1956) Results of the mountain home rainfall interception and infiltration project on black wattle, 1953–1954. Journal of South African Forestry 27: 72--85. Bernard JM (1963) Forest floor moisture capacity of the New Jersey pine barrens. Ecology 44(3): 574--576. Bouten W, Swart PJF, and De Water E (1991) Microwave transmission, a new tool in forest hydrological research. Journal of Hydrology 124(1–2): 119--130. Brechtel HM (1969) Wald und Abfluss-Methoden zur Erforschung der Bedeutung des Waldes fur das Wasserdargebot. Deutsche Gewasserkundliche Mitteilungen 8: 24--31. Breuer L, Eckhardt K, and Frede H-G (2003) Plant parameter values for models in temperate climates. Ecological Modelling 169(2–3): 237--293. Bruijnzeel LA (2005) Tropical montane cloud forest: A unique hydrological case. In: Bonell M and Bruijnzeel LA (eds.) Forests, Water and People in the Humid Tropics, pp. 462--483. Cambridge: Cambridge University Press. Bruijnzeel LA, Eugster W, and Burkard R (2005) Fog as a hydrologic input. In: Anderson MG (ed.) Encyclopedia of Hydrological Sciences, pp. 559--582. Chichester: Wiley. Bruijnzeel LA and Wiersum KF (1987) Rainfall interception by a young Acacia Auriculiformis (a. cunn) plantation forest in West Java, Indonesia: Application of Gash’s analytical model. Hydrological Processes 1: 309--319. Bryant ML, Bhat S, and Jacobs JM (2005) Measurements and modeling of throughfall variability for five forest communities in the southeastern US. Journal of Hydrology 312: 95--108. Calder IR (1986) A stochastic model of rainfall interception. Journal of Hydrology 89: 65--71. Calder IR (1990) Evaporation in the Uplands. Chichester: Wiley. Calder IR and Rosier PTW (1976) The design of large plastic-sheet net-rainfall gauges. Journal of Hydrology 30(4): 403--405.

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Calder IR and Wright IR (1986) Gamma ray attenuation studies of interception from Sitka Spruce: Some evidence for an additional transport mechanism. Water Resources Research 22: 409--417. Calder IR, Wright IR, and Murdiyarso D (1986) A study of evaporation from tropical rain forest–West Java. Journal of Hydrology 89: 13--31. Clark OR (1940) Interception of rainfall by prairie grasses, weeds and certain crop plants. Ecological Monographs 10: 243--277. Cuartas LA, Tomasella J, Nobre AD, Hodnett MG, Waterloo MJ, and Mnera JC (2007) Interception water-partitioning dynamics for a pristine rainforest in Central Amazonia: Marked differences between normal and dry years. Agricultural and Forest Meteorology 145(1–2): 69--83. de Groen MM and Savenije HHG (2006) A monthly interception equation based on the statistical characteristics of daily rainfall. Water Resources Research 42: W12417. Edwards WRN (1986) Precision weighing lysimetry for trees, using a simplified taredbalance design. Tree Physiology 1: 127--144. Fenicia F, Savenije HHG, Matgen P, and Pfister L (2008) Understanding catchment behavior through stepwise model concept improvement. Water Resources Research 44: 1--13. Friesen J, Beek C van, Selker J, Savenije HHG, and Giesen N van de (2008) Tree rainfall interception measured by stem compression. Water Resources Research 44: W00D15. Fritschen LJ, Cox L, and Kinerson R (1973) A 28-meter Douglas-fir in a weighing lysimeter. Forest Science 19: 256--261. Gash JHC (1979) An analytical model of rainfall interception by forests. Quarterly Journal of the Royal Meteorological Society 105: 43--55. Gash JHC, Lloyd CR, and Lauchaud G (1995) Estimation sparse forest rainfall interception with an analytical model. Journal of Hydrology 170: 79--86. Gash JHC and Morton AJ (1978) An application of the rutter model to the estimation of the interception loss from Thetford Forest. Journal of Hydrology 38(1–2): 49--58. Gash JHC, Rosier PTW, and Ragab R (2008) A note on estimating urban roof runoff with a forest evaporation model. Hydrological Processes 22(8): 1230--1233. Gash JHC and Stewart JB (1975) The average surface resistance of a pine forest derived from Bowen ratio measurements. Boundary-Layer Meteorology 8: 453--464. Gash JHC, Wright IR, and Lloyd CR (1980) Comparative estimates of interception loss from three coniferous forests in Great Britain. Journal of Hydrology 48(1–2): 89--105. Germer S, Elsenbeer H, and Moraes JM (2006) Throughfall and temporal trends of rainfall redistribution in an open tropical rainforest, South-Western Amazonia (Rondonia, Brazil). Hydrology and Earth System Sciences 10: 383--393. Gerrits AMJ, Pfister L, and Savenije HHG (2009a) Spatial and temporal variability of canopy and forest floor interception in a beech forest. Hydrological Processes, doi: 10.1002/hyp. 7712, published online, 7 june 2010. Gerrits AMJ, Savenije HHG, Hoffmann L, and Pfister L (2007) New technique to measure forest floor interception–an application in a beech forest in Luxembourg. Hydrology and Earth System Sciences 11: 695--701. Gerrits AMJ, Savenije HHG, and Pfister L (2009b) Canopy and forest floor interception and transpiration measurements in a mountainous beech forest in Luxembourg. IAHS Redbook 326: 18--24. Gerrits AMJ, Savenije HHG, Veling EJM, and Pfister L (2009c) Analytical derivation of the Budyko curve based on rainfall characteristics and a simple evaporation model. Water Resources Research 45: W04403. Grimmond CSB and Oke TR (1991) An evapotranspiration-interception model for urban areas. Water Resources Research 27: 1739--1755. Guevara-Escobar A, Gonzalez-Sosa E, Ramos-Salinas M, and Hernandez-Delgado GD (2007) Experimental analysis of drainage and water storage of litter layers. Hydrology and Earth System Sciences 11(5): 1703--1716. Hancock NH and Crowther JM (1979) A technique for the direct measurement of water storage on a forest canopy. Journal of Hydrology 41: 105--122. Haynes JL (1940) Ground rainfall under vegetation canopy of crops. Journal of the American Society of Agronomy 32: 176--184. Helvey JD (1964) Rainfall interception by hardwood forest litter in the southern Appalachians U.S. Forest Service Research Paper SE, vol. 8, pp. 1--8. Asherille, NC: Department of Agriculture, Forest Science, Southeastern Forest Experiment station. Helvey JD and Patric JH (1965) Canopy and litter interception of rainfall by Hardwoods of Eastern United States. Water Resources Research 1(2): 193--206. Herbst M, Rosier PT, McNeil DD, Harding RJ, and Gowing DJ (2008) Seasonal variability of interception evaporation from the canopy of a mixed deciduous forest. Agricultural and Forest Meteorology 148(11): 1655--1667.

Hildebrandt A, Al Aufi M, Amerjeed M, Shammas M, and Eltahir EAB (2007) Ecohydrology of a seasonal cloud forest in Dhofar: 1. Field experiment. Water Resources Research 43: W10411. Ho¨lscher D, Ko¨hler L, Dijk AIJM van, and Bruijnzeel LAS (2004) The importance of epiphytes to total rainfall interception by a tropical montane rain forest in Costa Rica. Journal of Hydrology 292(1–4): 308--322. Hoover MD and Lunt HA (1952) A key for the classification of forest humus types. Soil Science Society Proceedings 16: 368--371. Ho¨rmann G, Branding A, Clemen T, Herbst M, Hinrichs A, and Thamm F (1996) Calculation and simulation of wind controlled canopy interception of a beech forest in Northern Germany. Agricultural and Forest Meteorology 79(3): 131--148. Horton RE (1919) Rainfall interception. Monthly Weather Review 47(9): 603--623. Huang YS, Chen SS, and Lin TP (2005) Continuous monitoring of water loading of trees and canopy rainfall interception using the strain gauge method. Journal of Hydrology 311: 1--7. Hursh CR and Pereira HC (1953) Field moisture balance in the Shimba Hills, Kenya. East African Agricultural Journal 18: 139--148. Hutley LB, Doley D, Yates DJ, and Boonsaner A (1997) Water balance of an Australian subtropical rainforest at altitude: The ecological and physiological significance of intercepted cloud and fog. Australian Journal of Botany 45: 311--329. Keim R, Skaugset A, Link T, and Iroum A (2004) A stochastic model of throughfall for extreme events. Hydrology and Earth System Sciences 8(1): 23--34. Keim RF, Meerveld HJT van, and McDonnell JJ (2006a) A virtual experiment on the effects of evaporation and intensity smoothing by canopy interception on subsurface stormflow generation. Journal of Hydrology 327: 352--364. Keim RF, Skaugset AE, and Weiler M (2005) Temporal persistence of spatial patterns in throughfall. Journal of Hydrology 314: 263--274. Keim RF, Skaugset AE, and Weiler M (2006b) Storage of water on vergetation under simulated rainfall of varying intensity. Advances in Water Resources 29: 974--986. Kittredge J (ed.) (1948) Forest Influences. New York: McGraw-Hill. Klaassen W, Bosveld F, and de Water E (1998) Water storage and evaporation as constituents of rainfall interception. Journal of Hydrology 212–213: 36--50. Klaassen W, Lankreijer HJM, and Veen AWL (1996) Rainfall interception near a forest edge. Journal of Hydrology 185(1–4): 349--361. Li XY, Gong JD, Gao QZ, and Wei XH (2000) Rainfall interception loss by pebble mulch in the semi arid region of China. Journal of Hydrology 228: 165--173. Llorens P and Gallart F (2000) A simplified method for forest water storage capacity measurement. Journal of Hydrology 240: 131--144. Lloyd CR and Marques ADO (1988) Spatial variability of throughfall and stemflow measurements in amazonian rainforest. Agricultural and Forest Meteorology 42(1): 63--73. Lundberg A, Eriksson M, Halldin S, Kellner E, and Seibert J (1997) New approach to the measurement of interception evaporation. Journal of Atmospheric and Oceanic Technology 14: 1023--1035. Manfroi OJ, Kuraji K, Suzuki M, et al. (2006) Comparison of conventionally observed interception evaporation in a 100-m2 subplot with that estimated in a 4-ha area of the same Bornean Lowland tropical forest. Journal of Hydrology 329(1–2): 329--349. Miller HD (1966) Transport of intercepted snow from trees during snowstorms US Forest Service–Research Paper, vol. 33, pp. 1--30. Berkeley, CA: US department of Agriculture, Forest Service, Pacific Southwest Forest & Range Experiment Station. Miller JD, Anderson HA, Ferrier RC, and Walker TAB (1990) Comparison of the hydrological budgets and detailed hydrological responses in two forested catchments. Forestry 63(3): 251--269. Muzylo A, Llorens P, Valente F, Keizer J, Domingo F, and Gash J (2009) A review of rainfall interception modelling. Journal of Hydrology 370(1–4): 191--206. Nakayoshi M, Moriwaki R, Kawai T, and Kanda M (2009) Experimental study on rainfall interception over an outdoor urban-scale model. Water Resources Research 45: W04415. Navar J, Charles F, and Jurado E (1999) Spatial variations of interception loss components by Tamaulipan thornscrub in Northeastern Mexico. Forest Ecology and Management 24: 231--239. Oke TR (1982) The energetic basis of the urban heat island. Quarterly Journal of the Royal Meteorological Society 108(455): 1--24. Pathak PC, Pandey AN, and Singh JS (1985) Apportionment of rainfall in central Himalayan forests (India). Journal of Hydrology 76: 319--332. Penman HL (1948) Natural evaporation from open water, bare soil and grass. Proceedings of the Royal Society of London 193: 120--146. Pitman JI (1989) Rainfall interception by bracken litter–relationship between biomass, storage and drainage rate. Journal of Hydrology 111: 281--291. Putuhena W and Cordery I (1996) Estimation of interception capacity of the forest floor. Journal of Hydrology 180: 283--299.

Interception Pypker TG, Unsworth MH, and Bond BJ (2006) The role of epiphytes in rainfall interception by forests in the Pacific Northwest. I. Laboratory measurements of water storage. Canadian Journal of Forest Research 36: 808--818. Ragab R, Bromley J, Rosier P, Cooper JD, and Gash JHC (2003) Experimental study of water fluxes in a residential area: 1. Rainfall, roof runoff and evaporation: The effect of slope and aspect. Hydrological Processes 17(12): 2409--2422. Roberts MC and Klingeman PC (1970) The influence of landform and precipitation parameters on flood hydrograph. Journal of Hydrology 11: 393--411. Rowe L (1983) Rainfall interception by an evergreen beech forest, Nelson, New Zealand. Journal of Hydrology 66(1–4): 143--158. Rutter AJ, Kershaw KA, Robins PC, and Morton AJ (1971) A predictive model of rainfall interception in forests. I. Derivation of the model and comparison with observations in a plantation of Corsican pine. Agricultural Meteorology 9: 367--384. Rutter AJ, Morton AJ, and Robins PC (1975) A predictive model of rainfall interception in forests. II. Generalization of the model and comparison with observations in some coniferous and hardwood stands. Journal of Applied Ecology 12: 367--380. Sato Y, Kumagai T, Kume A, Otsuki K, and Ogawa S (2004) Experimental analysis of moisture dynamics of litter layers – the effect of rainfall conditions and leaf shapes. Hydrological Processes 18: 3007--3018. Savenije HHG (1997) Determination of evaporation from a catchment water balance at a monthly time scale. Hydrology and Earth System Sciences 1: 93--100. Savenije HHG (2004) The importance of interception and why we should delete the term evapotranspiration from our vocabulary. Hydrological Processes 18: 1507--1511. Savenije HHG (2005) Interception. In: Lehr JH and Keeley J (eds.) Water Encyclopedia: Surface and Agricultural Water. Hoboken, NJ: Wiley Publishers. Schaap MG and Bouten W (1997) Forest floor evaporation in a dense Douglas fir stand. Journal of Hydrology 193: 97--113. Shuttleworth WJ, Gash JHC, Lloyd CR, Moore CJ, Roberts JM, et al. (1984) Eddy correlation measurements of energy partition for Amazonian forest. Quarterly Journal of the Royal Meteorological Society 110: 1143--1162. Stewart JB (1977) Evaporation from the wet canopy of a pine forest. Water Resources Research 13(6): 915--921.

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Storck P, Lettenmaier DP, and Bolton SM (2002) Measurement of snow interception and canopy effects on snow accumulation and melt in a mountainous maritime climate, Oregon, United States. Water Resources Research 38: 1223. Thamm F and Widmoser P (1995) Zur hydrologischen Bedeutung der organischen Auflage im Wald: Untersuchungsmethoden und erste Ergebnisse. Zeitschrift fur Pflanzenernahrung und Bodenkunde 158: 287--292. Thurow TL, Blackburn WH, Warren SD, and Taylor CA Jr. (1987) Rainfall interception by midgrass, shortgrass, and live oak mottes. Journal of Range Management 40(5): 455--460. Toba T and Ohta T (2005) An observational study of the factors that influence interception loss in boreal and temperate forests. Journal of Hydrology 313: 208--220. Tobo´n-Marin C, Bouten IW, and Dekker S (2000) Gross rainfall and its partitioning into throughfall, stemflow and evaportation of intercepted water in four forest ecosystems in Western Amazonia. Journal of Hydrology 237: 40--57. Valente F, David JS, and Gash JHC (1997) Modelling interception loss for two sparse eucalypt and pine forest in central Portugal using reformulated Rutter and Gash analytical models. Journal of Hydrology 190: 141--162. Viville D, Biron P, Granier A, Dambrine E, and Probst A (1993) Interception in a mountainous declining spruce stand in the Strengbach catchment (Vosges, France). Journal of Hydrology 144: 273--282. Walsh RPD and Voigt PJ (1977) Vegetation litter: An underestimated variable in hydrology and geomorphology. Journal of Biogeography 4: 253--274. Wang A, Diao Y, Pei T, Jin C, and Zhu J (2007) A semi-theoretical model of canopy rainfall interception for a broad-leaved tree. Hydrological Processes 21(18): 2458--2463. Ward AD and Trimble SW (2004) Environmental hydrology, 2nd edn. Boca Raton, FL: CRC Press. Zhang GP and Savenije HHG (2005) Rainfall-runoff modelling in a catchment with a complex groundwater flow system: Application of the Representative Elementary Watershed (REW) approach. Hydrology and Earth System Sciences 9: 243--261. Ziegler AD, Giambelluca TW, Nullet MA, et al. (2009) Throughfall in an evergreendominated forest stand in Northern Thailand: Comparison of mobile and stationary methods. Agricultural and Forest Meteorology 149(2): 373--384. Zinke PJ (1967) Forest interception studies in the United States. In: Sopper WE and Lull HW (eds.) Forest Hydrology, pp. 137--161. Oxford: Pergamon.

2.05 Infiltration and Unsaturated Zone JW Hopmans, University of California, Davis, CA, USA & 2011 Elsevier B.V. All rights reserved.

2.05.1 Introduction 2.05.2 Soil Properties and Unsaturated Water Flow 2.05.2.1 Soil Water Retention 2.05.2.2 Unsaturated Hydraulic Conductivity 2.05.2.3 Modeling of Unsaturated Water Flow and Transport 2.05.2.4 Infiltration Processes 2.05.3 Infiltration Equations 2.05.3.1 Philip Infiltration Equation 2.05.3.2 Parlange et al. Model 2.05.3.3 Swartzendruber Model 2.05.3.4 Empirical Infiltration Equations 2.05.4 Measurements 2.05.4.1 Infiltration 2.05.4.2 Unsaturated Water Flow 2.05.5 Scaling and Spatial Variability Considerations 2.05.6 Summary and Conclusions Acknowledgments References

2.05.1 Introduction As soils make up the upper part of the unsaturated zone, they are subjected to fluctuations in water and chemical content by infiltration and leaching, water uptake by plant roots, and evaporation from the soil surface. It is the most dynamic region of the subsurface, as changes occur at increasingly smaller time and spatial scales when moving from the groundwater toward the soil surface. Environmental scientists are becoming increasingly aware that soils make up a critically important component of the earth’s biosphere, because of their food production and ecological functions, and the soil’s important role in controlling water quality. For example, prevention or remediation of soil and groundwater contamination starts with proper management of the unsaturated zone. Water entry into the soil by infiltration is among the most important soil hydrological processes, as it controls the partitioning between runoff and soil water storage. Runoff water determines surface water quantity and quality, whereas infiltrated water determines plant available water, evapotranspiration, groundwater recharge, and groundwater quality. Also through exfiltration, infiltrated water affects water quality in waterways and associated riparian zones. Despite its relevance and our reliable physical understanding of infiltration, we have generally many difficulties predicting infiltration at any scale. Mostly, this is so because the infiltration rate is a time-varying parameter of which its magnitude is largely controlled by spatially variable soil properties, in both vertical and horizontal directions of a hydrologic basin. Moreover, infiltration rate and runoff are affected by vegetation cover, as it protects the soil surface from the energy impacts of falling raindrops or intercepting rainfall, serving as temporary water storage. The kinetic energy of rainfall causes soil degradation, leading to soil surface sealing and decreasing infiltration.

103 103 104 105 105 106 108 108 109 109 110 110 110 111 111 112 113 113

Historically, solutions to infiltration problems have been presented by way of analytical solutions or empirically. Analytical solutions provide values of infiltration rate or cumulative infiltration as a function of time, making simplifying assumptions of soil depth variations of water content, before and during infiltration. Instead, we now often use powerful computers to conduct numerical simulations of unsaturated water flow to solve for water content and water fluxes throughout the unsaturated soil domain in a single vertical direction or in multiple spatial dimensions, allowing complex initial and boundary conditions. However, although the modeling of multidimensional unsaturated water flow is extremely useful for many vadose zone applications, it does not necessarily improve the soil surface infiltration rate prediction, in light of the large uncertainty of the soil physical properties and initial and boundary conditions that control infiltration. In contrast, empirical infiltration models serve primarily to fit model parameters to measured infiltration, but have limited power as a predictive tool.

2.05.2 Soil Properties and Unsaturated Water Flow The soil consists of a complex arrangement of mostly connected solid, liquid, and gaseous phases, with the spatial distribution and geometrical arrangement of each phase, and the partitioning of solutes between phases, controlled by physical, chemical, and biological processes. The unsaturated zone is bounded by the soil surface and merges with the groundwater in the capillary fringe. Water in the unsaturated soil matrix is held by capillary and adsorptive forces. Water is a primary factor leading to soil formation from the weathering of parent material such as rock or transported deposits, with additional factors of climate, vegetation, topography, and parent material determining soil physical properties.

103

104

Infiltration and Unsaturated Zone

Defining the soil’s dry bulk density by rb (M L3), soil porosity, e (L3 L3), is defined by

e¼1

rb rs

ð1Þ

with rs being the soil’s particle density (M L3). Equation (1) shows that soil porosity has lower values as bulk soil density is increased such as by compaction. Unsaturated water flow is largely controlled by the physical arrangement of soil particles in relation to the water and air phases within the soil’s pore space, as determined by pore-size distribution and water-filled porosity or volumetric water content, y (L3 water/L3 bulk soil). The volumetric water content y expresses the volume of water present per unit bulk soil as

y¼

wrb rw

ð2Þ

where w is defined as the mass water content (M of water/M dry soil) and we take rw ¼ 1000 kg m3. Alternatively, the soil water content can be described by the degree of saturation S (–) and the equivalent depth of stored water De (L), or

S¼

y e

and

De ¼ yDsoil

ð3Þ

so that y can also be defined by the equivalent depth of water per unit depth of bulk soil, Dsoil (L). The volumetric water content ranges between 0.0 (dry soil) and the saturated water content, ys, which is equal to the porosity if the soil were completely saturated. The degree of saturation varies between 0.0 (completely dry) and 1.0 (all pores completely waterfilled). When considering water flow, the porosity term is replaced by the saturated water content, ys, and both terms in Equation (3) are corrected by subtracting the so-called residual water content, yr (soil water content for which water is considered immobile), so that the effective saturation, Se, is defined as

Se ¼

y  yr ys  yr

weight of water, leading to soil water potential expressed by the equivalent height of a column of water (L). The resulting pressure head equivalent of the combined adsorptive and capillary forces in soils is defined as the matric pressure head, h. When expressed relative to the reference potential of free water, the water potential in unsaturated soils is negative (the soil water potential is less than the water potential of water at atmospheric pressure). Hence, the matric potential decreases or is more negative as the soil water content decreases. In using head units for water potential, the total water potential (H) is defined as the sum of matric potential (h), gravitational potential (z), hydrostatic pressure potential (p), and osmotic potential (p). For most hydrological applications, the contribution of the osmotic potential can be ignored, so that for unsaturated water flow (p ¼ 0) the total soil water potential can be written as

H ¼hþz

ð5Þ

The measurement of the soil water matric potential in situ is difficult and is usually done by tensiometers in the range of matric head values larger (less negative) than  6.0 m. A tensiometer consists of a porous cup, usually ceramic, connected to a water-filled tube (Young and Sisson, 2002). The suction forces of the unsaturated soil draw water from the tensiometer into the soil until the water pressure inside the cup (at pressure smaller than atmospheric pressure) is equal to the pressure equivalent of the soil water matric potential just outside the cup. The water pressure in the tensiometer is usually measured by a vacuum gauge or pressure transducer. Other devices that are used to indirectly measure the soil water matric potential include buried porous units (Scanlon et al., 2002), for which either the electrical resistance or the thermal conductivity is measured in situ, after coming into hydraulic equilibrium with the surrounding soil (h in sensor and soil are equal). Although widely used, these types of sensors require laboratory calibration, before field installation.

2.05.2.1 Soil Water Retention ð4Þ

In addition to the traditional thermogravimetric method to determine soil water content, many other measurement techniques are available, including neutron thermalization, electrical conductivity, dielectric, and heat pulse methods. A recent review on soil moisture measurement methods was presented by Robinson et al. (2008), focusing on measurement constraints between the many available methods across spatial scales. In soils, the driving force for water to flow is the gradient in total water potential. The total potential of bulk soil water can be written as the sum of all possible component potentials, so that the total water potential (ct) is equal to the sum of osmotic, matric, gravitational, and hydrostatic pressure potential. Whereas in physical chemistry the chemical potential of water is usually defined on a molar or mass basis, soil water potential is usually expressed with respect to a unit volume of water, thereby attaining units of pressure (Pa); or per unit

The soil water retention function determines the relation between the volume of water retained by the soil, expressed by y, and the governing soil matric, or suction forces (Dane and Hopmans, 2002). These suction forces are typically expressed by the soil water matric head (strictly negative) or soil suction (strictly positive). These suction forces increase as the size of the water-filled pores decreases, as may occur by drainage, water uptake by plant roots, or soil evaporation. Also known as the soil water release or soil water characteristic function, this soil hydraulic property describes the increase of y and the size of the water-filled pores with an increase in matric potential, as occurs by infiltration. Since the matric forces are controlled by pore-size distribution, specific surface area, and type of physico-chemical interactions at the solid–liquid interfaces, the soil water retention curve is very soil specific and highly nonlinear. It provides an estimate of the soil’s capacity to hold water after irrigation and free drainage (field capacity), minimum soil water content available to the plant (wilting point), and root zone water availability for plants.

Infiltration and Unsaturated Zone

The soil water retention curve exhibits hysteresis, that is, the y value is different for wetting (infiltration) and drying (drainage). By way of the unique relationship between soil water matric head and the radius of curvature of the air–water interface in the soil pores, and using the analogy between capillary tubes and the irregular pores in porous media, a relationship can be derived between soil water matric head (h) and effective pore radius, re, or

rgh ¼

2s cos a re

ð6Þ

where s and a are defined as the surface tension and wetting angle of wetting fluid with soil particle surface (typically values for s and a are 0.072 N m1 and 01, respectively ), r is the density of water, and g is the acceleration due to gravity (9.8 m s2). Because of capillary equation, the effective pore-size distribution can be determined from the soil water retention curve in the region where matric forces dominate. Laboratory and field techniques to measure the soil water retention curve, and functional models to fit the measured soil water retention data, such as the van Genuchten (1980) and Brooks and Corey (1964) models, are described by Kosugi et al. (2002). Alternatively, knowledge of the particle size distribution may provide information on the shape of the soil water retention curve, as presented by Nasta et al. (2009). An example of measured and fitted soil water retention data for two different soils is presented in Figure 1 (Tuli and Hopmans, 2004).

2.05.2.2 Unsaturated Hydraulic Conductivity The relation between the soil’s unsaturated hydraulic conductivity, K, and volumetric water content, y, is the second essential fundamental soil hydraulic property needed to

Soil matric potential head (cm)

describe unsaturated soil water flow. K is a function of the water and soil matrix properties, and controls water infiltration and drainage rates, and is strongly affected by water content and possibly by hysteresis. It is defined by the Darcy– Buckingham equation, which relates the soil water flux density to the total driving force for flow, with K being the proportionality factor. Except for special circumstances, the total driving force for water flow in soils is determined by the sum of the matric and gravitational forces, expressed by the total water potential head gradient, DH/L (L L1), where DH denotes the change in total water potential head over the distance L. For vertical flow, the application of Darcy’s law yields the magnitude of water flux from

 q ¼ KðyÞ

Measured Oso Flaco sand Optimized Oso Flaco sand Measured Columbia sandy loam Optimized Columbia sandy loam

 dh þ1 dz

ð7Þ

where q is the Darcy water flux density (L3 water L2 soil surface T1) and z defines the vertical position (z40, upwards, L). A soil system is usually defined by the bulk soil, without consideration of the size and geometry of the individual flow channels or pores. Therefore, the hydraulic conductivity (K) describes the ability of the bulk soil to transmit water, and is expressed by volume of water flowing per unit area of bulk soil per unit time (L T1). Functional models for unsaturated hydraulic conductivity are based on pore-size distribution, pore geometry, and connectivity, and require integration of soil water retention functions to obtain analytical expressions for the unsaturated hydraulic conductivity. The resulting expressions relate the relative hydraulic conductivity, Kr, defined as the ratio of the unsaturated hydraulic conductivity, K, and the saturated hydraulic conductivity, Ks, to the effective saturation, Se, and can be written in the following generalized form (Kosugi et al., 2002):

2Z

10 000

105

Se

3g

jhj Z dSe 7 6 6 7 Kr ðSe Þ ¼ Sle 6 Z0 S 7 4 5 Z jhj dSe

ð8Þ

0

1000

100

10

1 0

0.2

0.4

0.6

0.8

Sew Figure 1 Measured (symbols) and fitted (lines) soil water retention data. From Tuli AM and JW Hopmans (2004) Effect of degree of saturation on transport coefficients in disturbed soils. European Journal of Soil Science 55: 147–164.

where l and Z are parameters related to the tortuosity and connectivity of the soil pores, and the value of the parameter g is determined by the method of evaluating the effective pore radii. For values of l ¼ 0.5, Z ¼ 1.0, and g ¼ 2.0, Equation (8) reduces to the so-called Mualem (1976) model, that is routinely combined with the van Genuchten (1980) soil water retention model to yield a closed-form expression for the unsaturated hydraulic conductivity function. The moisture dependency is highly nonlinear, with a change in K of five or more orders of magnitude across field-representative changes in unsaturated soil water content. Methods to measure the saturation dependency of the hydraulic conductivity are involved and time consuming. A variety of methods are described in Dane and Topp (2002) and Dirksen (2001). Measurement errors are generally large due to (1) the difficulty of flow measurements in the low water content range and (2) the dominant effect of large pores (macropores), cracks, and fissures in the high water content range. An example of the unsaturated hydraulic conductivity for water, relative to its

Infiltration and Unsaturated Zone 1.0

1.0

0.8

0.8

0.6

0.6 K ra

K rw

106

0.4

0.4

0.2

0.2

0 (a)

0 0

0.2

0.4

0.6

0.8

1.0

(b)

0

0.2

0.4

0.6

0.8

1.0

Volumetric water content Oso Flaco fine sand

Columbia sandy loam

Measured Krw

Measured Krw

Measured Kra

Measured Kra

Measured Drg

Measured Drg

Measured ECra

Measured ECra

Figure 2 (a) Measured relative hydraulic conductivity for water (Krw) and (b) air conductivity (Kra) as a function of degree of water (Sew) and air (Sea) saturation. From Tuli AM and JW Hopmans (2004) Effect of degree of saturation on transport coefficients in disturbed soils. European Journal of Soil Science 55: 147–164.

saturated values (Krw), is presented in Figure 2(a), for the same two soils as in Figure 1. We note that ys in the vadose zone is typically about 85% of the porosity, so that a saturated soil (e.g., as the result of ponded infiltration) is really a satiated soil due to entrapped air, with a saturated hydraulic conductivity that is significantly smaller than the true Ks. The unsaturated hydraulic conductivity is related to the intrinsic soil permeability, k (L2), by

K¼

rgk m

ð9Þ

where m denotes the dynamic viscosity of water (F T L2). The usage of permeability instead of conductivity allows application of the flow equation to liquids other than water with different density and viscosity values. In addition to unsaturated hydraulic conductivity, Figure 2 also includes data for the saturation dependency of the relative air conductivity (Kra), as might be important for water infiltration in soil, when the soil gas phase is trapped and increasing in pressure, so that water infiltration is partly controlled by soil air permeability (Latifi et al., 1994).

2.05.2.3 Modeling of Unsaturated Water Flow and Transport Numerous studies have been published addressing different issues in the numerical modeling of unsaturated water flow using the Richards’ equation. In short, the dynamic water flow equation is a combination of the Darcy expression and a mass balance formulation. Using various solution algorithms, the soil region of interest is discretized in finite-size elements, i, that can be one, two, or three dimensional, to solve for temporal changes in h, y, or water flux, q, for each element or voxel i at any time t.

Most multidimensional soil water flow models use a finiteelement, Picard time-iterative numerical scheme (Sˇimunek et al., 2008) to solve the Richards equation. For isotropic conditions and one-dimensional vertical flow, the general water flow equation simplifies to

   qy q qh ¼ KðhÞ þ 1  Sðz; tÞ qt qz qz

ð10Þ

where S (L3 L3 T1) is the sink term, accounting for root water uptake. Boundary and initial conditions must be included to allow for specified soil water potentials or fluxes at all boundaries of the soil domain. Richards’ equation is a highly nonlinear partial differential equation, and is therefore extremely difficult to solve numerically because of the largely nonlinear dependencies of both water content and unsaturated hydraulic conductivity on the soil water matric head. Both the soil water retention and unsaturated hydraulic conductivity relationships must be known a priori to solve the unsaturated water flow equation. Specifically, it will need the slope of the soil water retention curve, or water capacity C(h), defined as CðhÞ ¼ dy=dh. As dissolved solutes move through the soils with the water, various physical, chemical, and biological soil properties control their fate. In addition to diffusion and dispersion, fate and transport of chemicals in the subsurface are influenced by sorption to the solid phase and biological transformations. Both diffusion and dispersion of the transported chemical are a function of pore-size distribution and water content. Mechanical or hydrodynamic dispersion is the result of water mixing within and between pores as a result of variations in pore water velocity. Increasing dispersivity values cause greater spreading of the chemical, thereby decreasing peak

Infiltration and Unsaturated Zone

2.05.2.4 Infiltration Processes For one-dimensional infiltration, the infiltration rate (L T1), i(t), can be defined by Equation (7) at the soil surface (subscript surf), or

 iðtÞ ¼ KðyÞ

 qh þ1 qz surf

ð11aÞ

Cumulative infiltration I(t), expressed as volume of water per unit soil surface area (L), is defined by

IðtÞ ¼

Z

t

iðtÞdt

ð11bÞ

0

Analytical solutions of infiltration generally assume that the wetted soil profile is homogeneous in texture with uniform initial water content. They also make distinction between ponded (h4 0 or p) and nonponded soil surface (unsaturated, ho0) infiltration. The infiltration capacity of the soil is defined by ic(t), the maximum rate at which a soil can absorb water for ponded soil surface conditions. Its maximal value is at time zero, and decreases with time to its minimum value approaching the soil’s saturated hydraulic conductivity, Ks, as the total water potential gradient decreases, and tends to unity, with the downward moving wetting front. As defined by Equation (11b), the soil’s cumulative infiltration capacity, Ic(t), is defined by the area under the capacity curve. It represents the maximum amount of water that the soil can absorb at any time. Typically, at the onset of infiltration (t ¼ 0), the rainfall rate, r(t), will be lower than ic(t), so that the infiltration rate is equal to the rainfall rate (i.e., r(t)oic(t) for hsurfo0). If at any point in time, the rainfall rate becomes larger than the infiltration capacity, ponding will occur (hsurf40), resulting in runoff. The time at which ponding occurs is defined as tp (time to ponding). Thus, the actual infiltration rate will depend on the rainfall rate and its temporal changes. This makes prediction of infiltration and runoff much more difficult for realistic time-variable rainfall patterns. Therefore, infiltration rate prediction is often described as a function of the cumulative infiltration, I, or i(I), independent of the time domain, and with i(I) curves that are independent of rainfall rate (Skaggs, 1982). An example of such a

time-invariant approach is the IDA or infiltrability-depth approximation (Smith et al., 2002). The main IDA assumption is that time periods between small rainfall events are sufficiently small so that soil water redistribution and evaporation between events do not affect infiltration rate. IDA implies that the infiltration rate at any given time depends only on the cumulative infiltration volume, regardless of the previous rainfall history. Following this approach, tp is defined as the time during a storm event when I becomes equal to Ic(tp), or

R¼

Z

tp

rðtÞdt ¼ Ic ðtp Þ

t¼0

whereas i(t) ¼ r(t) for totp. The time invariance of i(I) holds true also when a layered/sealed soil profile is considered (Mualem and Assouline, 1989). For illustration purposes, we present a hypothetical storm event with time-varying r(t) in Figure 3 (from Hopmans et al., 2007) in combination with an assumed soil-specific infiltration capacity curve, ic(t). At what time will ponding occur? It will not be at t ¼ 7, when r(t) exceeds ic for the first time. In order to approximate tp, we plot both Ic and R for the storm in Figure 4(a), as a function of time and determine tp as the time at which both curves intersect (tp ¼ 13, for R ¼ Ic ¼ 110), since at that time, the cumulative infiltration of the storm is identical to the soil’s infiltration capacity. The final corresponding i(I) for this soil and storm event is presented in Figure 4(b), showing that the soil infiltration rate is equal to r(t) until I ¼ R(t) ¼ Ic(tp) ¼ 110, after which the infiltration rate is soilcontrolled and determined by Ic(t). More accurate approximations to the time-invariant approach can be found in Sivapalan and Milly (1989) and Brutsaert (2005), using the time compression or time condensation approximation that more accurately estimates infiltration prior to surface ponding. In addition to whether the soil is ponded or not, solutions of infiltration distinguish between cases with and without gravity effects, as different analytical solutions apply. As Equation (11a) shows, infiltration rate i(t) is determined by both the soil water matric potential gradient, dh=dz, and gravity. However, at the early stage of infiltration into a relatively dry soil, infiltration rate is dominated by the matric potential gradient so that the gravity effects on infiltration can 15 Infiltration rate, i

concentration. Sorbed chemicals move through the vadose zone slower than noninteracting chemicals, and the degree of sorption will largely depend on mineral type, specific surface area of the solid phase, and organic matter fraction. In addition, biogeochemical processes and radioactive decay affect contaminant concentration, such as by cation exchange, mineral precipitation and dissolution, complexation, oxidation–reduction reactions, and by microbial biodegradation and transformations. However, all these mechanisms depend on soil environmental conditions, such as temperature, pH, water saturation, and redox status, and their soil spatial variations. The solute transport equation is generally referred to as the convection–dispersion equation (CDE), and includes the relevant transport mechanisms to simulate and predict temporal changes in soil solute concentration within the simulation domain (Sˇimunek et al., 2008).

107

r (t ) 10

5 ic 1 0

5

10

15

20

Time, t Figure 3 Hypothetical rainfall event, r(t), and soil infiltration capacity, ic(t). The rainfall event starts at t ¼ 0. From Hopmans JW, Assouline S, and Parlange J-Y (2007) Soil infiltration. In: Delleur JW (ed.) The Handbook of Groundwater Engineering, pp. 7.1–7.18. Boca Raton, FL: CRC Press.

108

Infiltration and Unsaturated Zone

Cumulative infiltration

160

Rainfall R (t ) Ic (t )

120 80 40

tp

0 0

5

10

20

Time

(a)

Infiltration rate, i

15

16 14 12 10 8 6 4 2 0

t p = 13

 I ¼ Ic ¼ K1 t þ ðhsurf  hf ÞDy ln 1 þ

ic (I )

0

50

100

150

Cumulative infiltration, I

(b)

Figure 4 (a) Cumulative infiltration corresponding with infiltration capacity, Ic(t), and cumulative rainfall, R(t) and (b) actual infiltration rate vs. I. Ponding starts only after tp ¼ 13, or I ¼ 110. From Hopmans et al. (2007).

be ignored. Gravity becomes important in the later stages of infiltration, when the wetting front has moved further down. For gravity-free drainage, a simple analytical solution can be found, after transforming Equation (11a) into a y-based form by defining the diffusivity DðyÞ ¼ KðyÞdh=dy, so that

iðtÞ ¼ DðyÞ

  qy q z surf

ð12Þ

Using the Boltzmann transformation for a constant head boundary condition (Bruce and Klute, 1956), and defining the scaling variable j ¼ z=t 1=2, combination of Equations (10) without gravity and sink term and (12) resulted in a unique solution of y as a function of j, from which the wetting profile can be computed for any time t (Kirkham and Powers, 1972). Defining y1 and y0 as the surface water content during infiltration and the initial uniform profile water content, respectively, cumulative infiltration, I, is computed from

I¼

Z

y1

z dy ¼ t 1=2

y0

Z

y1

j dy

ð13aÞ

y0

and results in the simple infiltration equation I ¼ St1/2, where the sorptivity S (L T1/2) is defined as

Sðy1 Þ ¼

Z

y1

j dy

horizontal infiltration, I is a linear function of t1/2, with S being defined as the slope of this line. Hence, for saturated soil conditions where y1 ¼ ys, the infiltration capacity is computed from ic(t) ¼ 12St1/2. Incidentally, this also leads to Ic ¼ S2/2ic. A relatively simple analytical solution without and with gravity effects was suggested by Green and Ampt (1911) for a ponded soil surface, with ysurf ¼ y1. The assumptions are that the wetting front can be approximated as a step function with a constant effective water potential, hf, at the wetting front, a wetting zone hydraulic conductivity of K(y1) ¼ K1 ¼ Ks, and a constant soil water profile of Dy ¼ y1  y0 . Using this so-called delta-function assumption of a D(y) with a Dirac-delta function form, both solutions for horizontal and vertical infiltration can be relatively easily obtained (Jury et al., 1991; Haverkamp et al., 2007). Assuming that K0 at the initial water content, y0, is negligible, the Green and Ampt (GA) solution of vertical infiltration for ponded conditions is (h ¼ hsurf40):

ð13bÞ

y0

Equation (13a) states that for gravity-free infiltration during the early times of vertical infiltration, and at all times for

ðhsurf

 I  hf ÞDy

ð14Þ

which can be solved iteratively for I. This simple, yet physically based, solution appears to work best for dry coarse-textured soils. A theoretical expression for the wetting front potential head,R hf, was defined by Mein and Farrell (1974), to yield that hf ¼ h00 Kr ðhÞdh; where the relative conductivity Kr ¼ K(h)/Ks. The so-called S-form of the GA equation can be obtained by comparing the gravity-free solution of GA with the Boltzmann solution, to yield S20 ¼ 2K1 hf Dy:

S21 ðy1 Þ ¼ 2K1 Dyðhsurf  hf Þ ¼ S20 þ 2K1 hsurf Dy

ð15aÞ

so that

  S20 I I ¼ K1 t þ hsurf Dy þ ln 1 þ hsurf Dy þ S20 =2K1 2K1

ð15bÞ

In reality, the wetting front is not a step function, but will consist of a time-dependent transition zone where water content changes from y1 to y0. The shape of this transition zone will be a function of time and is controlled by soil type. The step function assumption is better for uniform coarsetextured soils that have a Dirac-like D(y), for which there is a sharp decline in K with a decrease in water content near saturation. The wetting front is generally much more diffuse for finer-textured soils that have a wide pore-size distribution. By now, it must be clear that infiltration and its temporal changes are a function of many different soil factors. In addition to rainfall intensity and duration and the soil physical factors, such as soil water retention and hydraulic conductivity, infiltration is controlled by the initial water content, surface sealing and crusting, soil layering, and the ionic composition of the infiltrated water (Kutilek and Nielsen, 1994; Assouline, 2004). For example, Vandervaere et al. (1998) applied the GA model to sealed soil profiles, by assuming that the wetting front potential decreases suddenly as it leaves the seal and enters the soil. This results in a discontinuous drop in the infiltration rate. Many relatively simple infiltration equations have been proposed and are successfully used to

Infiltration and Unsaturated Zone

characterize infiltration. This has been achieved despite that these equations apply for homogeneous soils only, in theory.

2.05.3 Infiltration Equations In addition to the solutions in Section 2.05.2, other physically based analytical solutions have been presented, using different assumptions allowing for a closed-form solution. These can potentially be used to predict infiltration from known soil hydraulic properties of homogeneous soils. However, in practice, this is difficult as soil physical characteristics near the soil surface are time dependent because of soil structural changes and their high spatial variability. Alternatively, various empirical infiltration models have been proposed that are very useful for describing measured infiltration data. A parameter sensitivity analysis of many of the presented infiltration models, analyzing the effects of measurement error, was given by Clausnitzer et al. (1998). This section presents the most frequently used infiltration models in both categories.

2.05.3.1 Philip Infiltration Equation Philip (1957a) presented an analytical infinite-series solution to the water-content-based form of Richards’ equation for the case of vertical infiltration:

  qy q qy ¼ DðyÞ þ KðyÞ qt qz qz

ð16Þ

For the boundary condition of hsurf ¼ 0 and y1 ¼ ys , the Philip (1957a) solution converged to the true solution for small and intermediate times, but failed for large times. In this case, an alternative solution was presented (Philip, 1957b). With additional assumptions regarding the physical nature of soil water properties, Philip (1987) proposed joining solutions that are applicable for all times. Philip (1957c) introduced a truncation of the small-time series solution that is a simple two-parameter model equation (PH model):

Ic ¼ At þ St 1=2

ð17aÞ

which should be accurate for all but very large t, and suitable for applied hydrological studies. The sorptivity S depends on several soil physical properties, including initial water content y0, and the hydraulic conductivity and soil water retention functions. S is equal to the expression defined in Equation (13b). Philip (1969) showed that A may take values between 0.38Ks and 0.66Ks. The physical interpretation of A is not straightforward; however, for long times when gravity is dominant and hsurf ¼ 0, one would expect A to be equal to Ks. Differentiation of Equation (17a) yields the infiltration rate, or

ic ¼ 1=2St 0:5 þ A

ð17bÞ

Using (17b) to express t as a function of ic and substituting in Equation (17a) yields I(i), or

I¼

S 2 ði  A=2Þ 2ði  AÞ 2

ð17cÞ

109

For positive pressure heads (hsurf), the correction of Equation (15a) to S can be applied. In many cases, values of S and A are obtained from curve fitting. We note that for gravity-free flow, the pH solution without the gravity term corresponds with the Boltzmann solution for horizontal flow in Equation (13).

2.05.3.2 Parlange et al. Model Parlange et al. (1982) proposed the following universal model (Parlange et al., model, PA model):

  3 2dK1 I þ d  1 exp 62K1 7 S2 S2 6 7 t¼ 2 I  ln 4 5 2 2K1 ð1  dÞ S d 2

ð18aÞ

assuming that K0 is small so that the DK in Parlange et al. (1982) is equal to K1. The value of the parameter d can be chosen to approach various closed-form solutions. For example, Equation (18a) reduces to the GA solution for d equal to zero. Its value is a function of K(y), and is defined by (Parlange et al., 1985):

d¼

1 ys  y0

Z

ys y0

Ks  KðyÞ dy Ks

ð18bÞ

An approximate value of d ¼ 0.85 was suggested by Parlange et al. (1982) for a range of soil types. After taking the time derivative of I, the following i(I)-relationship can be derived (Espinoza, 1999):

  1 2IdK1 i ¼ K1 þ dK1 1  exp S2

ð18cÞ

Because Equation (18) is based on integration of the watercontent-based form of Richards’ equation, its theoretical scope is limited to nonponded conditions. A generalization of Equation (18) to include ponded conditions without affecting the value of S was introduced by Parlange et al. (1985). Haverkamp et al. (1990) presented a modification of their model to include upward water flow by capillary rise. The resulting infiltration model contained six physical parameters, in addition to the interpolation parameter d (Haverkamp et al., 1990). Both the PA and the Haverkamp et al. (1990) model require an iterative procedure to predict I(t). Barry et al. (1995) presented an explicit approximation to the Haverkamp et al. (1990) model, retaining all six physical parameters (BA model):

S 2 þ 2K1 hsurf Dy I ¼ K1 t þ 2DK    6ð2t  Þ 0:5 2t    t þ 1  g  exp  6 þ ð2t  Þ 0:5 3     g 2t 8 2:5 exp  ½1  ð1  gÞ t  þ 1 þ t 3   t þð2g þ t  Þln 1 þ g

ð19aÞ

110

Infiltration and Unsaturated Zone

where

t ¼

Another simple empirical infiltration equation is the Kostiakov (1932) model (KO):

2tðDKÞ 2 ; S 2 þ 2K1 hsurf Dy

g¼

2K1 ðhsurf þ ha ÞDy S2 þ 2K1 hsurf Dy

ð19bÞ

and ha denotes the absolute value of the soil water pressure head at which the air phase becomes discontinuous upon wetting. By defining

B1 ¼ ðhsurf þ ha ÞDy and B2 ¼

2 S 2 þ 2K1 hsurf Dy

ð19cÞ

Equation (19a) can be expressed by only four fitting parameters K0, K1, B1, and B2. The Clausnitzer et al. (1998) study concluded that both the PA and BA models described infiltration equally well; however, the BA model, while most advanced, was not as well suited to serve as a fitting model due to nonuniqueness problems caused by the larger number of fitting parameters.

i ¼ at b

ð22Þ

Clearly, this equation will not fit infiltration data at long times, as it predicts zero infiltration rate as t-N. The value of a should be equal to the infiltration rate at t ¼ 1, and 0obo1. Mezencev (1948) proposed another infiltration model, and modified the KO model by including a linear term with a coefficient b1, so that b1-K1 for t-N provided 0ob3o1 and b240 (ME model):

I ¼ b1 t þ

b2 tð1  b3 Þ 1  b3

ð23Þ

Other models include the Soil Conservation Service (1972) method and the Holtan solution (Kutilek and Nielsen, 1994; Espinoza, 1999).

2.05.3.3 Swartzendruber Model Swartzendruber (1987) proposed an alternative series solution that is applicable and exact for all infiltration times, and also allows for surface ponding. Its starting point is similar to the GA approach; however, its derivation does not require a step function for the wetted soil profile. Its simplified form is a three-parameter infiltration equation (SW model):

I ¼ K1 t þ

S 1  expðA0 t 1=2 Þ A0

ð20Þ

where A0 is a fitting parameter of which its value depends on the surface water content, y1. As A0-0, it reduces to a form of the Philip (1957b) model with K1 as the coefficient of the linear term, and for which dI/dt approaches K1 as t-N. As for the GA model, the S-term can be corrected using Equation (15a) to account for ponded conditions.

2.05.3.4 Empirical Infiltration Equations For most of these types of infiltration equations, the fitting parameters do not have a physical meaning and are evaluated by fitting to experimental data only. However, in many cases, the specific form of the infiltration equation is physically intuitive. For example, the empirical infiltration equation by Horton (1940) is one the most widely used empirical infiltration equations. It considers infiltration as a natural exhaustion process, during which infiltration rate decreases exponentially with time from a finite initial value, ic|t ¼ 0 ¼ (a1 þ a2), to a final value, a1 ¼ K1. Accordingly, cumulative infiltration I (L) is predicted as a function of time t (HO model):

I ¼ a1 t þ

a2 ½1  expða3 tÞ a3

ð21Þ

with the soil parameter a340, representing the decay of infiltration rate with time. In Equation (21), a1 can be associated with the hydraulic conductivity (LT1) of the wetted soil portion, K1, for t-N.

2.05.4 Measurements 2.05.4.1 Infiltration Infiltration measurements can serve various purposes. In addition to characterizing infiltration, for example, to compare infiltration between different soil types, or to quantify macropore flow, it is often measured to estimate the relevant soil hydraulic parameters from the fitting of the infiltration data to a specific physically based infiltration model. This is generally known as inverse modeling. Infiltration is generally measured using one of three different methods: a sprinkler method, a ring infiltration method, or a permeameter method. The sprinkler method is mostly applied to determine time of ponding for different water application rates, whereas the ring infiltrometer method is used when the infiltration capacity is needed. The permeameter method provides a way to measure infiltration across a small range of h-values p0. A general review of all three methods was recently presented by Smettem and Smith (Smith et al., 2002), whereas a comparison of different infiltration devices using seven criteria was presented by Clothier (2001). Rainfall sprinklers or rainfall simulators are also sprinkler infiltrometers, but they are typically used to study runoff and soil erosion (e.g., Morin et al., 1967). They mimic the rainfall characteristics (e.g., kinetic energy) of natural storms, specifically the rainfall rate, rainfall droplet size distribution, and drop velocity. Most of these devices measure infiltration by subtracting runoff from applied water. Using a range of water application rates, infiltration measurements can be used to determine the i(I) curve for a specific soil type, with specific soil hydraulic properties such as Ks or S. Various design parameters for many developed rainfall simulators, specifically nozzle systems, were presented by Peterson and Bubenzer (1986). A portable and inexpensive simulator for infiltration measurements along hillslopes was developed by Battany and Grismer (2000). This low-pressure system used a hypodermic syringe needle system to form uniform droplets at rainfall intensities ranging from 20 to 90 mm h1.

Infiltration and Unsaturated Zone

Ring infiltrometers have historically been used to characterize soil infiltration by determining the infiltration capacity, ic. A ring is carefully inserted in the soil so that water can be ponded over a known area. Since a constant head is required, a constant water level is maintained either by manually adding water and using a measuring stick to maintain a constant depth of ponded water, by using a Mariotte system, or by a valve connected to a float that closes at a predetermined water level. Measurements are usually continued until the infiltration rate is essentially constant. Water seepage around the infiltrometer is prevented by compaction of the soil around and outside of the infiltrometer. Multidimensional water flow under the ring is minimized by pushing the ring deeper into the soil, or by including an outer buffer ring. In the latter case, the soil between the two concentric rings is ponded at the same depth as the inner ring, to minimize lateral flow directed radially outward. The deviation from the assumed one-dimensionality depends on ring insertion depth, ring diameter, measurement time and soil properties such as its hydraulic conductivity, and the presence of restricting soil layers. A sensitivity analysis on diverging flow of infiltrometers was presented by Bouwer (1986) and Wu et al. (1997). Permeameters are generally smaller than infiltrometers and allow easy control of the soil water pressure head at the soil surface. Generally, multidimensionality of flow must be taken into account, using Wooding’s (1968) equation for steady flow (QN, L3 T1) from a shallow, circular surface pond of free water, or

 QN ¼ Ks

pr20

4r0 þ a

 ð24aÞ

The first and second terms in parentheses denote the gravitational and capillary components of infiltration and a denotes the parameter in Gardner’s (1958) unsaturated hydraulic conductivity function:

KðhÞ ¼ Ks expðahÞ

ð24bÞ

In this model of the so-called Gardner soil, the macroscopic capillary length, lc, is equivalent to 1/a. The basic analysis for most permeameter methods relies on Wooding’s solution. An extensive review of the use of permeameters was presented by Clothier (2001), including the tension infiltrometers and disk permeameters, by which the soil water pressure at water entry is controlled by a bubble tower. Their use is relatively simple, and based on analytical solutions of steady-state water flow. The permeameter method is economical in water use and portable. The soil hydraulic properties (S and K), in an inverse way, can be inferred from measurements using (1) both shortand long-time observations, (2) disks with various radii, or (3) using multiple water pressure heads. Transient solutions of infiltration may be preferable, as it allows analysis of shorter infiltration times, so that the method is faster and likely will better satisfy the homogeneous soil assumption. Differences between one- and three-dimensional solutions for transient infiltration were analyzed by Haverkamp et al. (1994), Vandervaere et al. (2000), and Smith et al. (2002) from multidimensional numerical modeling analysis. These effects were reported to be small if gravity effects were included.

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Nowadays, permeameters are most often applied to estimate the soil’s hydraulic characteristics in an inverse way, by fitting infiltration data to analytical solutions. In many cases, auxiliary water content or matric potential data are required to yield unique solutions.

2.05.4.2 Unsaturated Water Flow Whereas infiltration measures are typically conducted along the soil surface only, measurement of unsaturated water flow requires installation of instruments and sensors below ground, thereby largely complicating measurement procedures and analysis. The simplest expression for unsaturated water flow estimation is the Darcy equation (7), but still requires the measurement of soil water content (y) or soil water matric potential (h) at various soil depths, and knowledge of the unsaturated hydraulic function, K(y), as expressed by Equation (8). Installation of soil moisture or potential sensors requires extreme care, because of issues of soil disturbance, inadequate soil sensor contact, and inherent soil heterogeneities. In addition, it is not always straightforward to determine installation depth of sensors, as it will depend on a priori knowledge of soil horizon differentiation. Inherently problematic is the fact that no soil water flux meters are available to accurately measure the unsaturated soil water flux q in Equation (7). A review by Gee et al. (2003) provides possible direct and indirect methods, but none of them are adequate because of problems with divergence of water flow near the flux measurement device. Recently, the heat pulse probe was developed (Kamai et al., 2008) for indirect measurement of soil water flux, but is limited to fluxes of 6 mm d1 or higher. Finally, very few routine measurements are available to determine the K(y) relationship. In fact, the lack of the unsaturated conductivity information is the most limiting factor of in situ application of the Darcy equation. Most promising is the application of inverse modeling for parameter estimation of the soil hydraulic functions, using both laboratory and field techniques (Hopmans et al., 2002b), which can be used in conjunction with in situ water content and soil water potential measurements to estimate temporal changes in depth distribution of soil water flux. Selected steady-state solutions are provided in Jury et al. (1991), but are only of limited use for real field conditions since soil water content and matric potential values change continuously. Most realistically, one must apply the transient unsaturated water flow (Equation (10)) that arises from combination of the Darcy equation with mass conservation. However, its solution also requires a priori knowledge of the soil water capacity, C, as determined from the slope of the soil water retention curve, and time measurements of y and h, at the various soil horizon interfaces and at the boundaries of the soil domain of interest, including at the soil profile bottom. Although certainly possible, relatively few of such field experiments are conducted routinely because they are time consuming and wrought with complications. However, in combination with inverse modeling, such field experiments can provide a wealth of information, including plant root water uptake dynamics, plant transpiration, and drainage rates (Vrugt et al., 2001). Therefore, large lysimeters with selected

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Infiltration and Unsaturated Zone

water content and soil water potential measurements may be very useful.

2.05.5 Scaling and Spatial Variability Considerations Soil hydrologists need to apply locally measured soil physical data to characterize flow and transport processes at large-scale heterogeneous vadose zones. For example, prediction of soil water dynamics, such as infiltration at the field scale, is usually derived from the measurement of soil hydraulic properties from laboratory cores, as collected from a limited number of sampling sites across large spatial extents. Soil parameters obtained from these small-scale measurements are subsequently included in numerical models with a grid or element size many times larger, with the numerical results extrapolated to predict large-scale flow and transport behavior. Because of the typical nonlinearity of soil physical properties, their use across spatial scales is inherently problematic. Specifically, the averaging of processes determined from discrete small-scale samples may not describe the true soil behavior involving larger spatial structures. Moreover, the dominant physical flow processes may vary between spatial scales. Considering that soil physical, chemical, and biological measurements are typically conducted for small measurement volumes and that the natural variability of soils is enormous, the main question asked is how small-scale measurements can provide information about large-scale flow and transport behavior. In their treatise of scale issues of vadose zone modeling, Hopmans et al. (2002a) offer a conceptual solution, considering the control of small-scale processes on larger-scale flow behavior. Hence, vadose zone properties are nonunique and scale dependent, resulting in effective properties that vary across spatial scales and merely serve as calibration parameters in simulation models. Therefore, their accurate prediction in heterogeneous materials can only be accomplished using scale-appropriate measurements, including those that measure at the landscape scale. In addition, infiltration measurements are typically conducted at measurement scales in the range of 0.2–1.0 m. This is relevant for irrigation purposes, especially for micro-irrigation applications. Yet, infiltration information is often needed for much larger spatial scales, at the pedon scale, hillslope scale, and watershed scale. Very little work has been done relating infiltration process to measurement or support scale. Exceptions are the studies by Sisson and Wierenga (1981) and Haws et al. (2004), who measured steady-state infiltration at three spatial scales, ranging from 5 to 127- cm-diameter infiltrometer rings. Their results showed that much of the larger-scale infiltration occurs through smaller-scale regions, and that the spatial variability of infiltration decreased as the measurement scale increased. Thus, in general, we find that the process of infiltration might vary with spatial scale, and that larger spatial scales are required to estimate representative infiltration characteristics across a typical landscape. Many field studies have dealt with the significant areal heterogeneity of soil hydraulic properties, and particularly that of the saturated hydraulic conductivity, Ks (Nielsen et al., 1973). The heterogeneity in Ks is recognized to have a major effect on unsaturated flow, leading to significant variation in

local infiltration. In general, accounting for areal heterogeneity leads to shorter ponding times and to a more gradual decrease of the infiltration flux with time (Smith and Hebbert, 1979; Sivapalan and Wood, 1986). To characterize spatial variable infiltration rates, Sharma et al. (1980) measured infiltration with a double-ring infiltrometer at 26 sites in a 9.6-ha watershed. The infiltration data were fitted to the PH infiltration Equation (17a), and fitting parameters S and A were scaled to express their spatial variability and to describe the ensembleaverage or composite infiltration curve of the watershed. A simpler but similar scaling technique for infiltration data was presented by Hopmans (1989), who measured transient infiltration at 50 sites along a 100-m transect. Data were fitted to both the PH and a modified KO model that includes an additional constant c as a second term in Equation (23). This paper showed that spatial variability of infiltration can be easily described by the probability density function of a single scaling parameter, to be used for applications in Monte Carlo simulation of watershed hydrology, as suggested for the first time by Peck et al. (1977). For application at the field scale, the so-called one-point method was presented by Shepard et al. (1993) to estimate furrow-average infiltration parameters of PH Equation (17a), across a furrow-irrigated agricultural field. They used the volume-balance principle from furrow advance time across the field, water inflow rate, and flow area measurements. For modeling surface hydrology, by subtracting the infiltration rate, i(t), from the rainfall rate, r(t), it is possible to estimate spatial and temporal distributions of rainfall excess or runoff. The influence of spatial heterogeneity in rainfall and soil variability on runoff production was studied by Sivapalan and Wood (1986) from an analytical solution of infiltration and making use of the IDA approximation. Statistical characteristics of ponding time and infiltration rate were presented for two cases, one with a spatially variable soil with a lognormal Ks distribution and uniform rainfall, and the other for a homogeneous soil with spatially variable rainfall. Among the various results, this study concluded that the ensemble infiltration approach is biased for spatially variable soils. Their results also showed that the cumulative distribution of ponding times or proportion of ponded area is an excellent way of analyzing mean areal infiltration. Moreover, the spatial correction of infiltration rate is time dependent and varies depending on the correlation lengths of rainfall and soil Ks. This study neglected the effects of surface water run-on, as caused by accumulated water upstream, running on to neighboring areas, thereby contributing locally to infiltration. A quantitative analysis of soil variability effects on watershed hydraulic response that included surface water interactions, such as run-on, was presented by Smith and Hebbert (1979), through analysis of the effects of deterministic changes of infiltration properties in the direction of surface water flow, using a kinematic watershed model. In a subsequent study by Woolhiser et al. (1996), it was clearly demonstrated that runoff hydrographs along a hillslope are significantly affected by spatial trends in the soil’s saturated hydraulic conductivity. We expect that important new information can be collected by linking this interactive modeling approach with remote sensing and geographical information system (GIS)